U nited States February 2006 Environm ental Protection 

Age ncy EPA
 EPA 600/R-05/004aF 

Air Quality Criteria for 

Ozone and Related 

Photochemical Oxidants 

Volume I of III

EPA 600/R-05/004aF 

February 2006 

Air Quality Criteria for Ozone and Related 

Photochemical Oxidants 

Volume I 

National Center for Environmental Assessment-RTP Office 

Office of Research and Development 

U.S. Environmental Protection Agency 

Research Triangle Park, NC 
DISCLAIMER 

This document has been reviewed in accordance with U.S. Environmental Protection 

Agency policy and approved for publication. Mention of trade names or commercial products 

does not constitute endorsement or recommendation for use. 

PREFACE 

National Ambient Air Quality Standards (NAAQS) are promulgated by the United States 

Environmental Protection Agency (EPA) to meet requirements set forth in Sections 108 and 109 

of the U.S. Clean Air Act (CAA). Sections 108 and 109 require the EPA Administrator (1) to 

list widespread air pollutants that reasonably may be expected to endanger public health or 

welfare; (2) to issue air quality criteria for them that assess the latest available scientific 

information on nature and effects of ambient exposure to them; (3) to set “primary” NAAQS to 

protect human health with adequate margin of safety and to set “secondary” NAAQS to protect 

against welfare effects (e.g., effects on vegetation, ecosystems, visibility, climate, manmade 

materials, etc); and (5) to periodically review and revise, as appropriate, the criteria and NAAQS 

for a given listed pollutant or class of pollutants. 

In 1971, the U.S. Environmental Protection Agency (EPA) promulgated National Ambient 

Air Quality Standards (NAAQS) to protect the public health and welfare from adverse effects of 

photochemical oxidants. The EPA promulgates the NAAQS on the basis of scientific 

information contained in air quality criteria issued under Section 108 of the Clean Air Act. 

Following the review of criteria as contained in the EPA document, Air Quality Criteria for 

Ozone and Other Photochemical Oxidants published in 1978, the chemical designation of the 

standards was changed from photochemical oxidants to ozone (O ) in 1979 and a 1-hour O 3 3

NAAQS was set. The 1978 document focused mainly on the air quality criteria for O and, to a 3 

lesser extent, on those for other photochemical oxidants (e.g., hydrogen peroxide and the 

peroxyacyl nitrates), as have subsequent revised versions of the document. 

To meet Clean Air Act requirements noted above for periodic review of criteria and 

NAAQS, the O criteria document, Air Quality Criteria for Ozone and Other Photochemical 3

Oxidants, was next revised and released in August 1986; and a supplement, Summary of Selected 

I-ii 
New Information on Effects of Ozone on Health and Vegetation, was issued in January 1992. 

These documents were the basis for a March 1993 decision by EPA that revision of the existing 

1-h NAAQS for O3 was not appropriate at that time. That decision, however, did not take into 

account newer scientific data that had become available after completion of the 1986 criteria 

document. Such literature was assessed in the next periodic revision of the O air quality criteria 3 

document (O AQCD) which has completed in 1996 and provided scientific bases supporting the 3

setting by EPA in 1997 of the current 8-h O NAAQS. 3 

The purpose of this revised air quality criteria document for O and related photochemical 3 

oxidants is to critically evaluate and assess the latest scientific information published since that 

assessed in the above 1996 O AQCD, with the main focus being on pertinent new information 3 

useful in evaluating health and environmental effects data associated with ambient air O3 

exposures. However, other scientific data are also discussed in order to provide a better 

understanding of the nature, sources, distribution, measurement, and concentrations of O and 3 

related photochemical oxidants and their precursors in the environment. The document mainly 

assesses pertinent literature published through 2004, but also includes assessment of a few 

additional important studies published or accepted for publication in 2005. 

A First External Review Draft of this O AQCD (dated January 2005) was released for 3 

public comment and was reviewed by the Clean Air Scientific Advisory Committee (CASAC) 

in May, 2005 to obtain. Public comments and CASAC recommendations were then taken into 

account in making revisions to the document for incorporation into a Second External Review 

Draft (dated August, 2005), which underwent further public comment and CASAC review at a 

December, 2005 public meeting. Public comments and CASAC advice derived from review of 

that Second External Review Draft were considered in making revisions incorporated into this 

final version of the document (dated February, 2006). Evaluations contained in the present 

document will be drawn on to provide inputs to associated O Staff Paper analyses prepared by 3 

EPA’s Office of Air Quality Planning and Standards (OAQPS) to pose options for consideration 

by the EPA Administrator with regard to proposal and, ultimately, promulgation of decisions on 

potential retention or revision, as appropriate, of the current O NAAQS. 3 

Preparation of this document was coordinated by staff of EPA’s National Center for 

Environmental Assessment in Research Triangle Park (NCEA-RTP). NCEA-RTP scientific 

staff, together with experts from other EPA/ORD laboratories and academia, contributed to 

I-iii 
3 

3

writing of document chapters. Earlier drafts of document materials were reviewed by non-EPA 

experts in peer consultation workshops held by EPA. The document describes the nature, 

sources, distribution, measurement, and concentrations of O in outdoor (ambient) and indoor 3 

environments. It also evaluates the latest data on human exposures to ambient O and 

consequent health effects in exposed human populations, to support decision making regarding 

the primary, health-related O NAAQS. Lastly, the document also evaluates ambient O 3 3 

environmental effects on vegetation and ecosystems, surface level solar UV radiation flux and 

global climate change, and man-made materials to support decision making on secondary 

O NAAQS. 

NCEA acknowledges the valuable contributions provided by authors, contributors, and 

reviewers and the diligence of its staff and contractors in the preparation of this document. 

I-iv 
Air Quality Criteria for Ozone and Related 

Photochemical Oxidants 

VOLUME I

Executive Summary .................................................... E-1


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


2.	 PHYSICS AND CHEMISTRY OF OZONE IN THE ATMOSPHERE . . . . . . . 2-1


3.	 ENVIRONMENTAL CONCENTRATIONS, PATTERNS, AND 


EXPOSURE ESTIMATES .......................................... 3-1


4.	 DOSIMETRY, SPECIES HOMOLOGY, SENSITIVITY, AND


ANIMAL-TO-HUMAN EXTRAPOLATION . . . . . . . . . . . . . . . . . . . . . . . . . . . 4-1


5.	 TOXICOLOGICAL EFFECTS OF OZONE AND RELATED


PHOTOCHEMICAL OXIDANTS IN LABORATORY ANIMALS 


AND IN VITRO TEST SYSTEMS ................................... 5-1


6.	 CONTROLLED HUMAN EXPOSURE STUDIES OF OZONE AND 


RELATED PHOTOCHEMICAL OXIDANTS . . . . . . . . . . . . . . . . . . . . . . . . . . 6-1


7.	 EPIDEMIOLOGICAL STUDIES OF HUMAN HEALTH EFFECTS


ASSOCIATED WITH AMBIENT OZONE EXPOSURE . . . . . . . . . . . . . . . . . . 7-1


8.	 INTEGRATIVE SYNTHESIS: EXPOSURE AND HEALTH EFFECTS . . . . . 8-1


9.	 ENVIRONMENTAL EFFECTS: OZONE EFFECTS ON 


VEGETATION AND ECOSYSTEMS ................................ 9-1


10.	 TROPOSPHERIC OZONE EFFECTS ON UV-B FLUX AND


CLIMATE CHANGE PROCESSES ................................. 10-1


11.	 OZONE EFFECTS ON MAN-MADE MATERIALS . . . . . . . . . . . . . . . . . . . . 11-1


I-v 
Air Quality Criteria for Ozone and Related 

Photochemical Oxidants 

(cont’d) 

VOLUME II 

CHAPTER 2 ANNEX (ATMOSPHERIC PHYSICS/CHEMISTRY) . . . . . . . . . . AX2-1 

CHAPTER 3 ANNEX (AIR QUALITY AND EXPOSURE) . . . . . . . . . . . . . . . . . AX3-1 

CHAPTER 4 ANNEX (DOSIMETRY) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . AX4-1 

CHAPTER 5 ANNEX (ANIMAL TOXICOLOGY) . . . . . . . . . . . . . . . . . . . . . . . AX5-1 

CHAPTER 6 ANNEX (CONTROLLED HUMAN EXPOSURE) . . . . . . . . . . . . . AX6-1 

CHAPTER 7 ANNEX (EPIDEMIOLOGY) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . AX7-1 

VOLUME III 

CHAPTER 9 ANNEX (ENVIRONMENTAL EFFECTS) . . . . . . . . . . . . . . . . . . . AX9-1 

I-vi 
Table of Contents 

Page 

List of Tables . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . I-xviii


List of Figures . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . I-xxi


Authors, Contributors, and Reviewers . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . I-xxviii


U.S. Environmental Protection Agency Project Team for Development of Air

 Quality Criteria for Ozone and Related Photochemical Oxidants . . . . . . . . . . . . . . . . . . . I-xxxix 

U.S. Environmental Protection Agency Science Advisory Board (SAB) Staff Office

 Clean Air Scientific Advisory Committee (CASAC) Ozone Review Panel . . . . . . . . . . . . . I-xlii 

ABBREVIATIONS AND ACRONYMS . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . I-xlv 

EXECUTIVE SUMMARY ...................................................... E-1


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


1.1	 LEGAL AND HISTORICAL BACKGROUND . . . . . . . . . . . . . . . . . . . . . . . . . 1-1


1.1.1 	 Legislative Requirements ................................. 1-1


1.1.2	 Criteria and NAAQS Review Process . . . . . . . . . . . . . . . . . . . . . . . . 1-3


1.1.3 	 Regulatory Chronology ................................... 1-4


1.2	 CURRENT OZONE CRITERIA AND NAAQS REVIEW . . . . . . . . . . . . . . . . 1-8


1.2.1	 Key Milestones and Procedures for Document Preparation . . . . . . . 1-8


1.3	 ORGANIZATIONAL STRUCTURE OF THE DOCUMENT . . . . . . . . . . . . . 1-11


1.3.1 	 General Document Format ............................... 1-11


1.3.2 Organization and Content of the Document . . . . . . . . . . . . . . . . . . 1-12


REFERENCES ......................................................... 1-14


2.	 PHYSICS AND CHEMISTRY OF OZONE IN THE ATMOSPHERE . . . . . . . . . . . . . . 2-1


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


2.2	 CHEMICAL PROCESSES INVOLVED IN OZONE FORMATION 


AND DESTRUCTION ............................................ 2-2


2.3	 METEOROLOGICAL PROCESSES AFFECTING OZONE . . . . . . . . . . . . . . 2-8


2.4	 RELATIONS OF OZONE TO ITS PRECURSORS . . . . . . . . . . . . . . . . . . . . . 2-15


2.5	 THE ROLE OF CHEMISTRY-TRANSPORT MODELS IN 


UNDERSTANDING ATMOSPHERIC OZONE . . . . . . . . . . . . . . . . . . . . . . . 2-18


2.6	 TECHNIQUES FOR MEASURING OZONE AND ITS PRECURSORS . . . . 2-22


2.7 SUMMARY ................................................... 2-24


REFERENCES ......................................................... 2-27


3.	 ENVIRONMENTAL CONCENTRATIONS, PATTERNS, AND 


 EXPOSURE ESTIMATES ................................................ 3-1


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


3.2	 AMBIENT AIR QUALITY DATA FOR OZONE . . . . . . . . . . . . . . . . . . . . . . . 3-3


3.3	 SPATIAL VARIABILITY OF OZONE IN URBAN AREAS . . . . . . . . . . . . . 3-11


3.3.1	 Small-Scale Horizontal and Spatial Variability in 


Ozone Concentrations ................................... 3-14


I-vii 
Table of Contents 

(cont’d) 

Page 

3.4	 DIURNAL AND SEASONAL VARIABILITY OF OZONE . . . . . . . . . . . . . . 3-17


3.5	 TRENDS IN OZONE CONCENTRATIONS . . . . . . . . . . . . . . . . . . . . . . . . . . 3-32


3.6	 RELATIONSHIPS BETWEEN OZONE AND OTHER SPECIES . . . . . . . . . 3-39


3.7	 POLICY RELEVANT BACKGROUND OZONE CONCENTRATIONS . . . 3-44


3.8	 OZONE EXPOSURE IN VARIOUS MICROENVIRONMENTS . . . . . . . . . . 3-55


3.9 SUMMARY OF KEY POINTS .................................... 3-76


REFERENCES ......................................................... 3-80


4.	 DOSIMETRY, SPECIES HOMOLOGY, SENSITIVITY, AND ANIMAL-TO-


HUMAN EXTRAPOLATION .............................................. 4-1


4.1 	 INTRODUCTION ................................................ 4-1


4.2	 DOSIMETRY OF OZONE IN THE RESPIRATORY TRACT . . . . . . . . . . . . . 4-2


4.2.1	 Bolus-Response Studies in Humans . . . . . . . . . . . . . . . . . . . . . . . . . 4-4


4.2.2 	 General Uptake Studies ................................... 4-7


4.2.3 	 Dosimetry Modeling .................................... 4-10


4.2.4	 Summary and Conclusions - Dosimetry . . . . . . . . . . . . . . . . . . . . . 4-13


4.3	 SPECIES HOMOLOGY, SENSITIVITY, AND ANIMAL-TO-


HUMAN EXTRAPOLATION ..................................... 4-15


4.3.1	 Summary and Conclusions: Species Homology, Sensitivity, 


and Animal-to-Human Extrapolation . . . . . . . . . . . . . . . . . . . . . . . 4-22


REFERENCES ......................................................... 4-23


5.	 TOXICOLOGICAL EFFECTS OF OZONE AND RELATED PHOTOCHEMICAL


OXIDANTS IN LABORATORY ANIMALS AND IN VITRO TEST SYSTEMS . . . . . 5-1


5.1 	 INTRODUCTION ................................................ 5-1


5.2	 RESPIRATORY TRACT EFFECTS OF OZONE . . . . . . . . . . . . . . . . . . . . . . . 5-2


5.2.1 	 Biochemical Effects ...................................... 5-2


5.2.1.1	 Cellular Targets of Ozone Interaction . . . . . . . . . . . . . 5-2


5.2.1.2 	 Monooxygenases ............................. 5-5


5.2.1.3	 Antioxidants, Antioxidant Metabolism, and 


Mitochondrial Oxygen Consumption . . . . . . . . . . . . . . 5-5


5.2.1.4	 Lipid Metabolism and Content of the Lung . . . . . . . . . 5-8


5.2.1.5	 Ozone Interactions with Proteins and Effects 


on Protein Synthesis .......................... 5-11


5.2.1.6	 Differential Gene Expression . . . . . . . . . . . . . . . . . . . 5-12


5.2.1.7	 Summary and Conclusions—Biochemical Effects . . . 5-13


5.2.2 	 Lung Host Defenses ..................................... 5-14


5.2.2.1 	 Clearance .................................. 5-14


5.2.2.2	 Alveolar Macrophages . . . . . . . . . . . . . . . . . . . . . . . . 5-15


5.2.2.3 	 Immune System ............................. 5-17


5.2.2.4	 Interactions with Infectious Microorganisms . . . . . . . 5-20


5.2.2.5	 Summary and Conclusions—Lung Host Defenses . . . 5-21


I-viii 
Table of Contents 

(cont’d) 

Page 

5.2.3 Inflammation and Lung Permeability Changes . . . . . . . . . . . . . . . . 5-22


5.2.3.1	 Time Course of Inflammation and Lung 


Permeability Changes ......................... 5-23


5.2.3.2	 Concentration and Time of Exposure . . . . . . . . . . . . . 5-24


5.2.3.3	 Susceptibility Factors . . . . . . . . . . . . . . . . . . . . . . . . . 5-25


5.2.3.4	 Mediators of Inflammatory Response and Injury . . . . 5-28


5.2.3.5	 The Role of Nitric Oxide Synthase and Reactive 


Nitrogen in Inflammation . . . . . . . . . . . . . . . . . . . . . . 5-30


5.2.3.6	 Summary and Conclusions—Inflammation


and Permeability Changes . . . . . . . . . . . . . . . . . . . . . 5-32


5.2.4 Morphological Effects ................................... 5-35


5.2.4.1	 Acute and Subchronic Exposure Effects . . . . . . . . . . 5-35


5.2.4.2	 Summary of Acute and Subchronic 


Morphological Effects ........................ 5-39


5.2.4.3	 Subchronic and Chronic Exposure Effects . . . . . . . . . 5-41


5.2.4.4	 Summary and Conclusions—Subchronic and 


Chronic Morphological Effects . . . . . . . . . . . . . . . . . 5-44


5.2.5 Effects on Pulmonary Function . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5-45


5.2.5.1	 Acute and Subchronic Exposure Effects on 


Pulmonary Function .......................... 5-45


5.2.5.2	 Summary and Conclusions—Acute and 


Subchronic Effects on Pulmonary Function . . . . . . . . 5-47


5.2.5.3	 Ozone Effects on Airway Responsiveness . . . . . . . . . 5-47


5.2.5.4	 Summary and Conclusions—Effects on 


Airway Responsiveness ....................... 5-55


5.2.6 Genotoxicity Potential of Ozone . . . . . . . . . . . . . . . . . . . . . . . . . . . 5-56


5.2.6.1	 Summary and Conclusions—Genotoxicity 


Potential of Ozone ........................... 5-57


5.3	 SYSTEMIC EFFECTS OF OZONE EXPOSURE . . . . . . . . . . . . . . . . . . . . . . 5-57


5.3.1 Neurobehavioral Effects ................................. 5-57


5.3.2 Neuroendocrine Effects .................................. 5-59


5.3.3 Cardiovascular Effects ................................... 5-59


5.3.4 Reproductive and Developmental Effects . . . . . . . . . . . . . . . . . . . . 5-61


5.3.5 Effects on the Liver, Spleen, and Thymus . . . . . . . . . . . . . . . . . . . . 5-63


5.3.6 Effects on Cutaneous and Ocular Tissues . . . . . . . . . . . . . . . . . . . . 5-63


5.3.7 Summary and Conclusions—Systemic Effects of Ozone . . . . . . . . 5-64


5.4	 INTERACTIONS OF OZONE WITH OTHER CO-OCCURRING 


POLLUTANTS ................................................. 5-65


5.4.1 Ozone and Nitrogen Oxides .............................. 5-66


5.4.2 Ozone and Other Copollutants . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5-67


5.4.3 Complex (Multicomponent) Mixtures Containing Ozone . . . . . . . 5-69


I-ix 
Table of Contents 

(cont’d) 

Page 

5.4.4	 Summary and Conclusions—Interactions of Ozone with 


Other Co-Occurring Pollutants ............................ 5-76


5.5	 EFFECTS OF OTHER PHOTOCHEMICAL OXIDANTS . . . . . . . . . . . . . . . 5-77


5.5.1	 Summary and Conclusions—Effects of Other


Photochemical Oxidants ................................. 5-79


REFERENCES ......................................................... 5-80


6.	 CONTROLLED HUMAN EXPOSURE STUDIES OF OZONE AND 


RELATED PHOTOCHEMICAL OXIDANTS ................................. 6-1


6.1 	 INTRODUCTION ................................................ 6-1


6.2	 PULMONARY FUNCTION EFFECTS OF OZONE EXPOSURE IN 


HEALTHY SUBJECTS ........................................... 6-3


6.2.1 	 Introduction ............................................ 6-3


6.2.2 	 Acute Exposure for Up to 2 h .............................. 6-4


6.2.3 	 Prolonged Ozone Exposures ............................... 6-6


6.2.3.1	 Effect of Exercise Ventilation Rate on FEV1


Response to 6.6 h Ozone Exposure . . . . . . . . . . . . . . . 6-6


6.2.3.2	 Exercise Ventilation Rate as a Function of 


Body/Lung Size on FEV Response to 6.6 h 
 1 

Ozone Exposure .............................. 6-7


6.2.3.3	 Comparison of 2 h IE to 6.6 h O Exposure 3


Effects on Pulmonary Function . . . . . . . . . . . . . . . . . . 6-7


6.2.4 	 Triangular Ozone Exposures ............................... 6-8


6.2.5	 Mechanisms of Pulmonary Function Responses . . . . . . . . . . . . . . . 6-10


6.2.5.1	 Pathophysiologic Mechanisms . . . . . . . . . . . . . . . . . . 6-12


6.2.5.2	 Mechanisms at a Cellular and Molecular Level . . . . . 6-14


6.3	 SUBJECTS WITH PREEXISTING DISEASE . . . . . . . . . . . . . . . . . . . . . . . . . 6-16


6.3.1	 Subjects with Chronic Obstructive Pulmonary Disease . . . . . . . . . 6-16


6.3.2 	 Subjects with Asthma ................................... 6-16


6.3.3	 Subjects with Allergic Rhinitis . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6-18


6.3.4	 Subjects with Cardiovascular Disease . . . . . . . . . . . . . . . . . . . . . . . 6-20


6.4	 INTERSUBJECT VARIABILITY AND REPRODUCIBILITY 


OF RESPONSE ................................................. 6-21


6.5	 FACTORS MODIFYING RESPONSIVENESS TO OZONE . . . . . . . . . . . . . 6-23


6.5.1 	 Influence of Age ....................................... 6-23


6.5.2	 Gender and Hormonal Influences . . . . . . . . . . . . . . . . . . . . . . . . . . 6-24


6.5.3	 Racial, Ethnic, and Socioeconomic Status Factors . . . . . . . . . . . . . 6-25


6.5.4	 Influence of Physical Activity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6-25


6.5.5 	 Environmental Factors ................................... 6-26


6.5.6	 Oxidant-Antioxidant Balance . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6-27


6.5.7 	 Genetic Factors ........................................ 6-28


6.6 	 REPEATED O EXPOSURE EFFECTS . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6-29
 3


I-x 
Table of Contents 

(cont’d) 

Page 

6.7 EFFECTS ON EXERCISE PERFORMANCE . . . . . . . . . . . . . . . . . . . . . . . . . 6-30


6.8 EFFECTS ON AIRWAY RESPONSIVENESS . . . . . . . . . . . . . . . . . . . . . . . . 6-30


6.9 EFFECTS ON INFLAMMATION AND HOST DEFENSE . . . . . . . . . . . . . . 6-32


6.9.1 	 Introduction ........................................... 6-32


6.9.2	 Inflammatory Responses in the Upper Respiratory Tract . . . . . . . . 6-32


6.9.3	 Inflammatory Response in the Lower Respiratory Tract . . . . . . . . . 6-33


6.9.4	 Adaptation of Inflammatory Responses . . . . . . . . . . . . . . . . . . . . . 6-38


6.9.5	 Effect of Anti-Inflammatory and Other Mitigating Agents . . . . . . . 6-39


6.9.6	 Changes in Host Defense Capability Following 


Ozone Exposures ....................................... 6-40


6.10 EXTRAPULMONARY EFFECTS OF OZONE . . . . . . . . . . . . . . . . . . . . . . . 6-42


6.11 EFFECTS OF OZONE MIXED WITH OTHER POLLUTANTS . . . . . . . . . . 6-43


6.12 CONTROLLED STUDIES OF AMBIENT AIR EXPOSURES . . . . . . . . . . . 6-44


6.12.1 	 Mobile Laboratory Studies ............................... 6-44


6.13 SUMMARY ................................................... 6-44


REFERENCES ......................................................... 6-47


7.	 EPIDEMIOLOGIC STUDIES OF HUMAN HEALTH EFFECTS ASSOCIATED 


WITH AMBIENT OZONE EXPOSURE ...................................... 7-1


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


7.1.1	 Approach to Identifying Ozone Epidemiologic Studies . . . . . . . . . . 7-2


7.1.2	 Approach to Assessing Epidemiologic Evidence . . . . . . . . . . . . . . . 7-3


7.1.3	 Considerations in the Interpretation of Epidemiologic Studies 


of Ozone Health Effects .................................. 7-5


7.1.3.1	 Exposure Assessment and Measurement Error


in Epidemiologic Studies ....................... 7-6


7.1.3.2	 Ozone Exposure Indices Used . . . . . . . . . . . . . . . . . . 7-10


7.1.3.3	 Lag Time: Period between Ozone Exposure 


and Observed Health Effect . . . . . . . . . . . . . . . . . . . . 7-11


7.1.3.4	 Model Specification to Adjust for Temporal 


Trends and Meteorologic Effects . . . . . . . . . . . . . . . . 7-14


7.1.3.5	 Confounding Effects of Copollutants . . . . . . . . . . . . . 7-17


7.1.3.6	 Hypothesis Testing and Model Selection in 


Ozone Epidemiologic Studies . . . . . . . . . . . . . . . . . . 7-18


7.1.3.7	 Impact of Generalized Additive Models 


Convergence Issue on Ozone Risk Estimates . . . . . . . 7-21


7.1.3.8	 Summary of Considerations in the Interpretation 


of Ozone Epidemiologic Studies . . . . . . . . . . . . . . . . 7-23


7.1.4	 Approach to Presenting Ozone Epidemiologic Evidence . . . . . . . . 7-25


7.2 FIELD STUDIES ADDRESSING ACUTE EFFECTS OF OZONE . . . . . . . . 7-26


7.2.1	 Summary of Key Findings on Field Studies of Acute 


Ozone Effects from the 1996 Ozone AQCD . . . . . . . . . . . . . . . . . . 7-26


I-xi 
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7.2.2	 Introduction to Recent Field Studies of Acute Ozone Effects . . . . . 7-27


7.2.3	 Effects of Acute Ozone Exposure on Lung Function . . . . . . . . . . . 7-28


7.2.3.1 Spirometry (FEV ) Studies in Outdoor Worker,
 1 

Exercise, Children, Elderly, and Asthmatic Panels ... 7-29


7.2.3.2	 Peak Flow Meter (PEF) Studies in Asthmatics 


and Healthy Individuals ....................... 7-40


7.2.4 	 Respiratory Symptoms ................................... 7-48


7.2.5 	 Acute Airway Inflammation .............................. 7-55


7.2.6	 Acute Ozone Exposure and School Absences . . . . . . . . . . . . . . . . . 7-57


7.2.7 	 Cardiovascular Endpoints ................................ 7-60


7.2.7.1	 Cardiac Autonomic Control . . . . . . . . . . . . . . . . . . . . 7-60


7.2.7.2	 Acute Myocardial Infarction . . . . . . . . . . . . . . . . . . . . 7-63


7.2.7.3	 Cardiovascular Endpoints in Human 


Clinical Studies .............................. 7-64


7.2.7.4	 Summary of Field Studies with Cardiovascular 


Outcomes .................................. 7-64


7.2.8	 Summary of Field Studies Assessing Acute Ozone Effects . . . . . . 7-65


7.3	 ACUTE EFFECTS OF OZONE ON DAILY EMERGENCY 


DEPARTMENT VISITS AND HOSPITAL ADMISSIONS . . . . . . . . . . . . . . 7-66


7.3.1	 Summary of Key Findings on Studies of Emergency 


Department Visits and Hospital Admissions from the 


1996 Ozone AQCD ..................................... 7-66


7.3.2	 Review of Recent Studies of Emergency Department Visits 


for Respiratory Diseases ................................. 7-66


7.3.3	 Studies of Hospital Admissions for Respiratory Diseases . . . . . . . 7-71


7.3.3.1	 All-year and Seasonal Effects of Ozone on 


Respiratory Hospitalizations . . . . . . . . . . . . . . . . . . . . 7-72


7.3.3.2	 Potential Confounding of the Ozone Effect on 


Respiratory Hospitalizations by Copollutants . . . . . . 7-79


7.3.4	 Association of Ozone with Hospital Admissions for 


Cardiovascular Disease .................................. 7-80


7.3.5	 Summary of Acute Ozone Effects on Daily Emergency 


Department Visits and Hospital Admissions . . . . . . . . . . . . . . . . . . 7-83


7.4	 ACUTE EFFECTS OF OZONE ON MORTALITY . . . . . . . . . . . . . . . . . . . . 7-83


7.4.1	 Summary of Key Findings on Acute Effects of Ozone on 


Mortality from the 1996 Ozone AQCD . . . . . . . . . . . . . . . . . . . . . . 7-83


7.4.2	 Introduction to Assessment of Current Ozone-


Mortality Studies ....................................... 7-84


7.4.3	 Single-Pollutant Model Ozone-Mortality Risk Estimates . . . . . . . . 7-85


7.4.4	 Meta-Analyses of Ozone-Mortality Risk Estimates . . . . . . . . . . . . 7-94


7.4.5	 Seasonal Variation in Ozone-Mortality Risk Estimates . . . . . . . . . 7-97


7.4.6	 Ozone-Mortality Risk Estimates Adjusting for PM Exposure . . . 7-100


I-xii 
Table of Contents 

(cont’d) 

Page 

7.4.7	 Ozone Risk Estimates for Specific Causes of Mortality . . . . . . . . 7-103


7.4.8	 Ozone-Mortality Risk Estimates for Specific Subpopulations . . . 7-108


7.4.9	 Summary of Acute Ozone Effects on Mortality . . . . . . . . . . . . . . 7-110


7.5	 EFFECTS OF CHRONIC OZONE EXPOSURE . . . . . . . . . . . . . . . . . . . . . . 7-111


7.5.1	 Summary of Key Findings on Studies of Health Effects and 


Chronic Ozone Exposure from the 1996 Ozone AQCD . . . . . . . . 7-111


7.5.2	 Introduction to Morbidity Effects of Chronic Ozone Exposure . . 7-111


7.5.3	 Seasonal Ozone Effects on Lung Function . . . . . . . . . . . . . . . . . . 7-113


7.5.4	 Chronic Ozone Exposure Effects on Lung Function and 


Respiratory Symptoms .................................. 7-115


7.5.5	 Chronic Ozone Exposure and Respiratory Inflammation . . . . . . . 7-122


7.5.6 	 Risk of Asthma Development ............................ 7-124


7.5.7	 Respiratory Effects of Chronic Ozone Exposure on 


Susceptible Populations ................................. 7-126


7.5.8	 Effects of Chronic Ozone Exposure on Mortality and 


Cancer Incidence ...................................... 7-127


7.5.9	 Effects of Ozone on Birth-Related Health Outcomes . . . . . . . . . . 7-131


7.5.10	 Summary of Chronic Ozone Exposure Effects on Morbidity 


and Mortality ......................................... 7-134


7.6	 INTERPRETIVE ASSESSMENT OF THE EVIDENCE IN 


EPIDEMIOLOGIC STUDIES OF OZONE HEALTH EFFECTS . . . . . . . . . 7-134


7.6.1 	 Introduction .......................................... 7-134


7.6.2 	 Ozone Exposure Indices ................................ 7-135


7.6.3	 Confounding by Temporal Trends and Meteorologic Effects


in Time-Series Studies .................................. 7-137


7.6.3.1	 Assessment of Ozone Effects after Adjusting for


Temporal Trends and Meteorologic Effects . . . . . . . 7-138


7.6.3.2	 Importance of Season-Specific Estimates of


Ozone Health Effects ........................ 7-141


7.6.4	 Assessment of Confounding by Copollutants . . . . . . . . . . . . . . . . 7-148


7.6.4.1	 Relationship between Personal Exposure to 


Ozone and Copollutants ...................... 7-149


7.6.4.2	 Assessment of Confounding Using 


Multipollutant Regression Models . . . . . . . . . . . . . . 7-150


7.6.5	 Concentration-Response Function and Threshold . . . . . . . . . . . . . 7-154


7.6.6	 Heterogeneity of Ozone Health Effects . . . . . . . . . . . . . . . . . . . . . 7-159


7.6.7	 Health Effects of Ozone in Susceptible and Vulnerable 


Populations .......................................... 7-163


7.6.7.1	 Health Effects Associated with Ambient 


Ozone Exposure in Asthmatics . . . . . . . . . . . . . . . . 7-164


7.6.7.2	 Age-Related Differences in Ozone Effects . . . . . . . . 7-168


I-xiii 
Table of Contents 

(cont’d) 

Page 

7.6.7.3	 Vulnerability of Outdoor Workers and Others


Who Participate in Outdoor Activities to Ozone 


Health Effects .............................. 7-171


7.6.8	 Summary of Key Findings and Conclusions Derived from 


Ozone Epidemiologic Studies ............................ 7-174


REFERENCES ........................................................ 7-178


8. INTEGRATIVE SYNTHESIS: OZONE EXPOSURE AND HEALTH EFFECTS . . . . . 8-1


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


8.1.1 	 Chapter Organization ..................................... 8-2


8.2	 AMBIENT OZONE AIR QUALITY IN THE UNITED STATES . . . . . . . . . . . 8-3


8.2.1	 Current Ozone Concentrations and Spatial Patterns . . . . . . . . . . . . . 8-3


8.2.2	 Diurnal and Seasonal Variations . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8-4


8.2.3 	 Long-Term Trends ....................................... 8-4


8.2.4	 Interrelationships Between Ozone and Other Ambient Pollutants . . 8-5


8.2.5	 Policy Relevant Background (PRB) Ozone Concentrations . . . . . . . 8-6


8.3	 FACTORS AFFECTING HUMAN EXPOSURE TO AMBIENT OZONE . . . . 8-8


8.3.1 	 Personal Exposure ....................................... 8-8


8.3.2 	 Indoor Concentrations .................................... 8-9


8.4	 SYNTHESIS OF AVAILABLE INFORMATION ON OZONE­


RELATED HEALTH EFFECTS ................................... 8-10


8.4.1	 Integration of Experimental and Epidemiologic Evidence . . . . . . . 8-12


8.4.1.1	 Cross-Cutting Issues Relevant to Assessment/


Interpretation of Ozone Health Effects . . . . . . . . . . . . 8-12


8.4.1.2 	 Dosimetry .................................. 8-14


8.4.2	 Experimental Evidence for Ozone-Related Health Effects . . . . . . . 8-15


8.4.3	 Biological Basis for O Health Effects Assessment . . . . . . . . . . . . 8-28
 3 

8.4.4 	 Epidemiologic Evidence ................................. 8-32


8.4.4.1	 Acute Ozone Exposure Studies . . . . . . . . . . . . . . . . . 8-33


8.4.4.2	 Chronic Ozone Exposure Studies . . . . . . . . . . . . . . . . 8-39


8.4.4.3	 Summary of the Epidemiologic Evidence . . . . . . . . . 8-40


8.5	 ASSESSMENT OF POTENTIAL THRESHOLDS . . . . . . . . . . . . . . . . . . . . . 8-42


8.6	 BIOLOGICAL PLAUSIBILITY AND COHERENCE OF EVIDENCE


FOR OZONE-RELATED HEALTH EFFECTS . . . . . . . . . . . . . . . . . . . . . . . . 8-44


8.6.1	 Acute Ozone Exposure-Induced Health Effects . . . . . . . . . . . . . . . 8-45


8.6.2	 Chronic O Exposure-Induced Health Effects . . . . . . . . . . . . . . . . . 8-50
 3 

8.6.3	 Mortality-Related Health Endpoints . . . . . . . . . . . . . . . . . . . . . . . . 8-51


8.6.4	 Health Effects of Ozone-Containing Pollutant Mixtures . . . . . . . . 8-53


8.7	 SUSCEPTIBLE AND VULNERABLE POPULATIONS, AND 


POTENTIAL PUBLIC HEALTH IMPACTS . . . . . . . . . . . . . . . . . . . . . . . . . . 8-55


8.7.1	 Preexisting Disease as a Potential Risk Factor . . . . . . . . . . . . . . . . 8-56


I-xiv 
Table of Contents 

(cont’d) 

Page 

8.7.2	 Age-Related Variations in Susceptibility/Vulnerability . . . . . . . . . 8-59


8.7.3	 Vulnerability of Outdoor Workers and Others Who 


Participate in Outdoor Activities . . . . . . . . . . . . . . . . . . . . . . . . . . . 8-61


8.7.4	 Genetic Factors Affecting O 3
Susceptibility . . . . . . . . . . . . . . . . . . 8-63

8.7.5	 Potential Public Health Impacts . . . . . . . . . . . . . . . . . . . . . . . . . . . 8-65


8.7.5.1	 Concepts Related to Defining of Adverse 


Health Effects .............................. 8-65


8.7.5.2	 Estimation of Potential Numbers of Persons 


in At-Risk Susceptible Population Groups in 


the United States ............................. 8-70


8.8 SUMMARY AND CONCLUSIONS FOR OZONE HEALTH EFFECTS . . . 8-73


REFERENCES ......................................................... 8-83


APPENDIX 8A:	Summary of New Animal Toxicology, Human Clinical, 

and U.S./Canadian Epidemiologic Studies of Health Effects 

Associated with Ambient or Near-Ambient Ozone Exposures . . . . . . 8A-1 

9.	 ENVIRONMENTAL EFFECTS: OZONE EFFECTS ON VEGETATION 


AND ECOSYSTEMS .................................................... 9-1


9.1 	 INTRODUCTION ................................................ 9-1


9.2	 METHODOLOGIES USED IN VEGETATION RESEARCH . . . . . . . . . . . . . 9-3


9.3	 SPECIES RESPONSE AND MODE-OF-ACTION . . . . . . . . . . . . . . . . . . . . . . 9-6


9.4	 MODIFICATION OF FUNCTIONAL AND GROWTH RESPONSES . . . . . . 9-7


9.5	 EFFECTS-BASED AIR QUALITY EXPOSURE INDICES . . . . . . . . . . . . . . 9-11


9.6	 OZONE EXPOSURE-PLANT RESPONSE RELATIONSHIPS . . . . . . . . . . . 9-15


9.7	 EFFECTS OF OZONE EXPOSURE ON NATURAL ECOSYSTEMS . . . . . 9-19


9.8 ECONOMICS .................................................. 9-22


REFERENCES ......................................................... 9-25


10.	 THE ROLE OF TROPOSPHERIC OZONE IN UVB-RELATED HUMAN 


HEALTH OUTCOMES AND IN CLIMATE CHANGE . . . . . . . . . . . . . . . . . . . . . . . . 10-1


10.1 	 INTRODUCTION ............................................... 10-1


10.2	 THE ROLE OF TROPOSPHERIC OZONE IN DETERMINING 


GROUND-LEVEL UV-B FLUX ................................... 10-1


10.2.1	 Factors Governing Ultraviolet Radiation Flux at the 


Earth’s Surface ........................................ 10-1


10.2.1.1	 UV Radiation: Wavelengths and Energies . . . . . . . . 10-2


10.2.1.2	 Temporal Variations in Solar Flux at the 


Earth’s Surface .............................. 10-4


10.2.1.3	 Atmospheric Radiative Interactions with 


Solar Ultraviolet Radiation . . . . . . . . . . . . . . . . . . . . . 10-5


I-xv 
Table of Contents 

(cont’d) 

Page 

10.2.1.4	 Data Requirements for a Surface 


UV-B Climatology .......................... 10-15


10.2.2	 Factors Governing Human Exposure to Ultraviolet Radiation . . . 10-16


10.2.2.1 	 Outdoor Activities .......................... 10-16


10.2.2.2 	 Occupation ................................ 10-18


10.2.2.3 	 Age ...................................... 10-18


10.2.2.4 	 Gender ................................... 10-19


10.2.2.5 	 Geography ................................. 10-20


10.2.2.6	 Protective Behavior . . . . . . . . . . . . . . . . . . . . . . . . . 10-20


10.2.2.7	 Summary of Factors that Affect Human 


Exposures to Ultraviolet Radiation . . . . . . . . . . . . . 10-21


10.2.3	 Factors Governing Human Health Effects due to 


Ultraviolet Radiation ................................... 10-21


10.2.3.1 	 Erythema .................................. 10-22


10.2.3.2 	 Skin Cancer ............................... 10-24


10.2.3.3	 Ultraviolet Radiation Exposure and the 


Incidence of Nonmelanoma Skin Cancers . . . . . . . . 10-25


10.2.3.4	 Ocular Effects of Ultraviolet Radiation 


Exposure .................................. 10-32


10.2.3.5	 Ultraviolet Radiation and Immune System 


Suppression ................................ 10-34


10.2.3.6	 Protective Effects of Ultraviolet Radiation—


Production of Vitamin D ..................... 10-35


10.2.4	 Summary and Conclusions for Ozone Effects on UV-B Flux . . . . 10-37


10.3 TROPOSPHERIC OZONE AND CLIMATE CHANGE . . . . . . . . . . . . . . . . 10-38


10.3.1	 The Projected Impacts of Global Climate Change . . . . . . . . . . . . 10-39


10.3.2	 Solar Energy Transformation and the Components of the 


Earth’s Climate System ................................. 10-43


10.3.3	 The Composition of the Atmosphere and the Earth’s 


Radiative Equilibrium ................................ 10-44

10.3.3.1	 Forcing of the Earth’s Radiative Balance . . . . . . . . . 10-45


10.3.4	 Factors Affecting the Magnitude of Climate Forcing 


by Ozone ............................................ 10-47


10.3.4.1	 The Global Burden of Tropospheric Ozone . . . . . . . 10-48


10.3.4.2	 Background Concentrations versus Regionally-


Oriented Ozone Enhancements . . . . . . . . . . . . . . . . 10-49


10.3.4.3	 Ozone Trends: Globally and in North America . . . . 10-51


10.3.4.4	 The Sensitivity of Ozone-Related Forcing 


Surface to Albedo ........................... 10-54


10.3.4.5	 The Altitude Dependence of Forcing by 


Tropospheric Ozone ......................... 10-54


I-xvi 
Table of Contents 

(cont’d) 

Page 

10.3.4.6 Co-occurrence of Ozone with Particulate Matter . . . 10-54


10.3.5	 Estimated Forcing by Tropospheric Ozone . . . . . . . . . . . . . . . . . . 10-55


10.3.5.1 Direct Climate Forcing Due to Ozone . . . . . . . . . . . 10-55


10.3.5.2 Indirect Forcing Due to Ozone . . . . . . . . . . . . . . . . . 10-57


10.3.5.3 Predictions for Future Climate Forcing by 


Anthropogenic Ozone ........................ 10-58


10.3.6	 The Impact of a Warming Climate on Atmospheric


Ozone Concentrations .................................. 10-59


10.3.7 Conclusion ........................................... 10-59


REFERENCES ........................................................ 10-61


11. OZONE EFFECTS ON MAN-MADE MATERIALS . . . . . . . . . . . . . . . . . . . . . . . . . . . 11-1


11.1 ELASTOMERS ................................................. 11-1


11.2 TEXTILES AND FABRICS ....................................... 11-3


11.3 DYES, PIGMENTS, AND INKS ................................... 11-4


11.4 ARTISTS’ PIGMENTS ........................................... 11-5


11.5 SURFACE COATINGS ......................................... 11-12


11.6 CONCLUSIONS ............................................... 11-13


REFERENCES ........................................................ 11-15


I-xvii 
List of Tables 

Number Page 

1-1 National Ambient Air Quality Standards (NAAQS) for Ozone . . . . . . . . . . . . . . . . . . 1-5 

1-2 Key Milestones for Development of Revised Ozone Air Quality 

Criteria Document ................................................... 1-10 

3-1 Summary Statistics for the Spatial Variability of O (in ppm) in Selected 3

Urban Areas in the United States ........................................ 3-13 

3-2 Previous Estimates of Background O in Surface Air Over the United States . . . . . . 3 3-48 

3-3 Personal Exposure Concentrations . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3-64 

3-4 Indoor/Outdoor Ozone Concentrations in Various Microenvironments . . . . . . . . . . 3-65 

5-1 Summary of Studies that Evaluated Morphological Effects of a Single 

Acute O Exposure 3 ................................................... 5-40 

7-1a Field Studies that Investigated the Association Between Acute Ambient O 3

Exposure and Changes in FEV in Adults 1 ................................. 7-30 

7-1b Percent Changes in FEV (95% CI) Associated with Acute Ambient O 1 3

Exposures in Adults, Ordered by Size of the Estimate . . . . . . . . . . . . . . . . . . . . . . . . 7-31 

7-1c Cross-day Percent Changes in FEV (95% CI) Associated with Acute 1

Ambient O Exposures in Adults, Ordered by Size of the Estimate . . . . . . . . . . . . . . 3 7-32 

7-2a Field Studies that Investigated the Association Between Acute Ambient O 3

Exposure and Changes in FEV in Children . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1 7-33 

7-2b Percent Changes in FEV (95% CI) Associated with Acute Ambient O 1 3

Exposures in Children, Ordered by Size of the Estimate . . . . . . . . . . . . . . . . . . . . . . 7-34 

7-2c Cross-day Percent Changes in FEV (95% CI) Associated with Acute 1

Ambient O Exposures in Children, Ordered by Size of the Estimate . . . . . . . . . . . . 3 7-35 

7-3 Difference in Annual Percent Increases in Lung Function from the Least 

to the Most Polluted Community in the Children’s Health Study by Time 

Spent Outdoors . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7-119 

I-xviii 
List of Tables 

(cont’d)


Number Page


8-1 Acute O -Induced Physiological and Biochemical Changes in Human 
 3 

and Animals ........................................................ 8-29


8-2	 Gradation of Individual Responses to Short-Term Ozone Exposure in 


Healthy Persons ..................................................... 8-67


8-3	 Gradation of Individual Responses to Short-Term Ozone Exposure in Persons 


with Impaired Respiratory Systems ...................................... 8-68


8-4	 Prevalence of Selected Cardiorespiratory Disorders by Age Group and


by Geographic Region in the United States (2002 [U.S. Adults] and 2003 

[U.S. Children] National Health Interview Survey) . . . . . . . . . . . . . . . . . . . . . . . . . . 8-71


8-5 Acute Respiratory Conditions per 100 Persons/Year by Age Group in the 


United States (1996 National Health Interview Survey) . . . . . . . . . . . . . . . . . . . . . . 8-72


8A-1 Short-Term Ozone-Induced Health Effects Observed in Controlled Human 


Exposure Studies . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8A-2 

8A-2 Effects of Acute O 3
Exposure on Lung Function in the U.S. and Canada . . . . . . . . . 8A-6

8A-3	 Effects of Acute O Exposure on Asthma Emergency Department Visits 3


in the U.S. and Canada ............................................... 8A-10 

8A-4 Effects of Acute O Exposure on Total Respiratory and Asthma Hospital 3


Admissions in the U.S. and Canada . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8A-13 

8A-5 Effects of Acute O 3
Exposure on All-Cause Mortality in the U.S. and Canada . . . . 8A-18

8A-6	 Toxicological Effects of Acute Ozone Exposure in Animals . . . . . . . . . . . . . . . . . . 8A-23 

10-1	 Examples of Impacts Resulting From Projected Changes in Extreme


Climate Events ..................................................... 10-41


10-2	 CTM Studies Assessed by the IPCC for its Estimate of the Change in 


Global and Total Column O Since the Preindustrial Era . . . . . . . . . . . . . . . . . . . . 10-52
 3 

10-3	 Tropospheric O Change (O ) in Dobson Units (DU) Since Preindustrial 3 3


Times, and the Accompanying Net (SW plus LW) Radiative Forcings !2 (Wm ), After Accounting for Stratospheric Temperature Adjustment 


(using the Fixed Dynamical Heating Method) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 10-56


I-xix 
List of Tables 

(cont’d) 

Number Page 

11-1 Average 24-h Ozone Concentrations Producing the Highest Frequency 

of Cracks of a Certain Length in the Middle and Central Zones of the 

Rubber Test Strips ................................................... 11-3 

11-2 Cuprammonium Fluidity of Moist Cotton Cloth Exposed to 20 to 60 ppb Ozone . . 11-4 

11-3 Color Change After 12 Weeks of Exposure to a Mixture of 

Photochemical Oxidants . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11-11 

I-xx 
List of Figures 

Number Page 

2-1 Schematic overview of O photochemistry in the stratosphere and troposphere 3 . . . . . 2-4 

2-2a Surface weather chart showing sea level (MSL) pressure (kPa), and 

surface fronts ........................................................ 2-9 

2-2b Vertical cross section along dashed line (a-aN) from northwest to the southeast . . . . . 2-9 

2-3 The diurnal evolution of the planetary boundary layer while high pressure 

prevails over land .................................................... 2-11 

2-4 Locations of low level jet occurrences in decreasing order of prevalence 

(most frequent, common, observed) ...................................... 2-12 

2-5 Conceptual two-reservoir model showing conditions in the PBL and in the 

3 lower free troposphere during a multiday O episode . . . . . . . . . . . . . . . . . . . . . . . . 2-13 

2-6 A scatter plot of daily maximum 8-h average O concentrations versus daily 3

maximum temperature for May through September 1994 to 2004 in the 

Baltimore, MD Air Quality Forecast Area . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2-14 

2-7 A scatter plot of daily maximum 8-h average O concentrations versus daily 3

maximum temperature for May through September 1996 to 2004 at sites 

downwind of Phoenix, AZ ............................................. 2-15 

2-8 Measured values of O and NO (NO – NO ) during the afternoon at rural 3 z y x

sites in the eastern United States (grey circles) and in urban areas and urban 

plumes associated with Nashville, TN (gray dashes); Paris, France (black 

diamonds); and Los Angeles CA (Xs) .................................... 2-18 

2-9 Main components of a comprehensive atmospheric chemistry modeling 

system, such as Models-3 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2-19 

3-1 Countywide mean daily maximum 8-h O concentrations, May to 3

September 2000 to 2004 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3-4 

3-2 Countywide 95th percentile value of daily maximum 8-h O concentrations, 3

May to September 2000 to 2004 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3-5 

3-3 Box plots showing daily maximum 8-h O averaged by month over 1993 to 3

2002 in the five regions in the eastern United States derived by Lehman 

et al. (2004) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3-7 

I-xxi 
List of Figures 

(cont’d)


Number Page


3-4a-c Hourly average O concentrations observed at selected (a) rural-agricultural 
 3 

(b) rural-forested, and (c) rural-residential or commercial sites for 2004 . . . . . . . . . . 3-8


3-5a-d Daily 8-h maximum O concentrations observed at selected national park sites . . . . 3-10
 3


3-6 Vertical profile of O obtained over low vegetation . . . . . . . . . . . . . . . . . . . . . . . . . . 3-16
 3


3-7 Vertical profile of O obtained in a spruce forest . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3-17
 3


3-8 Composite, nationwide diurnal variability in hourly averaged O in urban areas . . . 3-18
 3


3-9 Composite, nationwide diurnal variability in 8-h average O in urban areas . . . . . . . 3-19
 3


3-10a-f Diurnal variability in hourly averaged O in selected urban areas . . . . . . . . . . . . . . . 3-20
 3


3-10g-l Diurnal variability in hourly averaged O in selected urban areas . . . . . . . . . . . . . . . 3-21
 3


3-11a-f Diurnal variability in 8-h O in selected urban areas . . . . . . . . . . . . . . . . . . . . . . . . . 3-23
 3


3-11g-l Diurnal variability in 8-h O in selected urban areas . . . . . . . . . . . . . . . . . . . . . . . . . 3-24
 3


3-12a-d Diurnal variations in hourly averaged O on weekdays and weekends in 3


four cities .......................................................... 3-26


3-12e-h Diurnal variations in hourly averaged O on weekdays and weekends in 
 3


four cities .......................................................... 3-27


3-13a-d Diurnal variations in 8-h average O on weekdays and weekends in four cities . . . . 3-28
 3


3-13e-h Diurnal variations in 8-h average O on weekdays and weekends in four cities . . . . 3-29
 3


3-14a-f Diurnal variability in 8-h average O in selected urban areas . . . . . . . . . . . . . . . . . . . 3-30
 3


3-14g-l Diurnal variability in 8 hour averaged O in selected urban areas . . . . . . . . . . . . . . . 3-31
 3


3-15 Composite diurnal variability in hourly O concentrations observed at 3


CASTNET sites ..................................................... 3-33


3-16 Composite diurnal variability in 8-h O concentrations observed at 
 3


CASTNET sites ..................................................... 3-33


I-xxii 
List of Figures 

(cont’d) 

Number Page 

3-17 Year-to-year variability in nationwide mean daily maximum 8-h O 3

concentrations . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3-34 

3-18 Year-to-year variability in nationwide 95th percentile value of the daily 

maximum 8-h O3 concentrations ........................................ 3-35 

3-19a-h Year-to-year variability in mean daily maximum 8-h O concentrations at 3

selected national park (NP), national wildlife refuge (NWR), and national 

monument (NM) sites . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3-36 

3-20a-h Year-to-year variability in 95th percentile of daily maximum 8-h O 3

concentrations at selected national park (NP), national wildlife refuge 

(NWR), and national monument (NM) sites . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3-37 

3-21 Binned mean PM concentrations versus binned mean O concentrations 2.5 3

observed at Fort Meade, MD from July 1999 to July 2001 . . . . . . . . . . . . . . . . . . . . 3-40 

3-22 The co-occurrence pattern for O and nitrogen dioxide using 2001 data 3

from the AQS ....................................................... 3-43 

3-23 The co-occurrence pattern for O and sulfur dioxide using 2001 data from AQS 3 .... 3-44 

3-24 The co-occurrence pattern for O and PM using 2001 data from AQS . . . . . . . . . . 3 2.5 3-45 

3-25a Monthly maximum hourly average O concentrations at Yellowstone 3

National Park (WY) in 1998, 1999, 2000, and 2001 . . . . . . . . . . . . . . . . . . . . . . . . . 3-46 

3-25b Hourly average O concentrations at Yellowstone National Park (WY) 3

for the period January to December 2001 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3-46 

3-26 Estimates of background contribution to surface afternoon (13 to 17 LT) 

O concentrations in the United States as a function of local O concentration, 3 3

site altitude, and season ............................................... 3-50 

3-27 Time-series of hourly average O concentrations observed at five national 3

parks: Denali (AK), Voyageur (MN), Olympic (WA), Glacier (MT), and 

Yellowstone (WY) ................................................... 3-54 

3-28 Hypothetical exposure time profile: pollutant exposure as a function of time 

showing how the average exposure, integrated exposure, and peak exposure 

relate to the instantaneous exposure . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3-56 

I-xxiii 
List of Figures 

(cont’d)


Number Page


3-29 Conceptual overview of an exposure model . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3-60


4-1 Structure of lower airways with progression from the large airways to 


the alveolus .......................................................... 4-3


4-2 Ozone uptake fraction as a function of volumetric penetration (V ) in a 
 P 

representative subject .................................................. 4-5


4-3 Ozone uptake efficiency as a function of breathing frequency at a minute 


ventilation of 30 L/min ................................................. 4-9


5-1 Schematic overview of ozone interaction with epithelial lining fluid and 


lung cells ............................................................ 5-3


5-2 The major cellular targets and proposed mechanisms of ozone toxicity in 


the lung ............................................................. 5-6


5-3 Mechanisms of ozone toxicity ........................................... 5-9


5-4 Mouse chromosomes on which genes or gene loci have been identified that 


modulate responses to O3 .............................................. 5-33


5-5 Schematic comparison of the duration-response profiles for epithelial 


hyperplasia, bronchoalveolar exudation, and interstitial fibrosis in the 

centriacinar region of lung exposed to a constant low concentration of ozone . . . . . 5-36


6-1 Ozone-induced changes in FEV (top panel) and O concentration profiles 
 1 3


(bottom panel) as a function of exposure duration . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6-9


6-2 Recovery of FEV responses following a 2 h exposure to 0.4 ppm O with IE . . . . . 6-12
 1 3


6-3 Predicted O -induced decrements in FEV as a function of exposure duration 3 1


and level of IE (line labels are V levels) in young healthy adults (20 yrs of age) 0 E 

exposed to 0.3 ppm O3 ................................................ 6-26


6-4 Time course of acute responses seen in humans exposed to O3 . . . . . . . . . . . . . . . . . 6-35


7-1 Percent change (95% CI) in morning PEF in children per standardized 


increment .......................................................... 7-41


7-2 Percent change (95% CI) in afternoon PEF in children per standardized 


increment .......................................................... 7-42


I-xxiv 
List of Figures 

(cont’d)


Number Page


Percent changes in PEF per 30 ppb increase in 8-h avg O in urban children . . . . . . 7-44
 3


Percent change in PEF per 30 ppb increase in 8-h avg O with a cumulative 3


lag of 1 to 5 days . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7-45


Odds ratios for the incidence of cough among asthmatic children per 


standardized increment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7-49


Odds ratios for extra medication use among asthmatic children per 


standardized increment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7-50


Odds ratio for the incidence of symptoms per 30 ppb increase in 8-h avg O
3


with a cumulative lag of 1 to 4 days . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7-52


Ozone-associated percent change (95% CI) in emergency department visits 


for asthma per standardized increment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7-68


Ozone-associated percent change (95% CI) in total respiratory hospitalizations


for all-year analyses per standardized increment . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7-73


Ozone-associated percent change (95% CI) in total respiratory hospitalizations


by season per standardized increment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7-74


Percent changes in total respiratory hospitalizations per 40 ppb increase 

in 1-h max O in children less than two years of age during the summer 3


(May to August) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7-78


Ozone-associated percent change (95% CI) in total respiratory hospitalizations


with adjustment for PM indices per standardized increment . . . . . . . . . . . . . . . . . . . 7-80


Ozone-associated percent change (95% CI) in total cardiovascular


hospitalizations per standardized increment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7-81


All cause (nonaccidental) O excess mortality risk estimates (95% CI) for
 3


all-year analyses per standardized increment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7-86


All cause (nonaccidental) O excess mortality risk estimates (95% CI) for
 3


all-year analyses per standardized increment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7-87


Median 24-h avg O concentrations (10th percentile to 90th percentile range) 
 3


for 95 U.S. communities (NMMAPS) from 1987 to 2000, arranged by 


O3 concentration . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7-89


7-3


7-4


7-5


7-6


7-7


7-8


7-9


7-10


7-11


7-12


7-13


7-14


7-15


7-16


I-xxv 
List of Figures 

(cont’d) 

Number Page 

8-1A,B Frequency distributions of FEV changes following 6.6-h exposures to a 1

constant concentration of O3 or filtered air . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8-17 

8-2 Frequency distributions of FEV changes following 6.6-h exposures to a 1

constant concentration of O3 or filtered air . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8-19 

8-3 Resolution time-line for the respiratory, physiological, and biochemical 

parameters are derived from studies reported in Chapter 6 and Chapter 6 Annex ... 8-30 

8-4 Acute (1-8 h) O exposure-induced cellular and molecular changes and 3

timelines for their resolution depicted here are derived from the data 

reported in Leikauf et al. (1995) and Mudway and Kelly (2000) . . . . . . . . . . . . . . . . 8-31 

9-1 Common anthropogenic stressors and the essential ecological attributes 

they affect .......................................................... 9-20 

10-1 Complexity of factors that determine human exposure to UV radiation . . . . . . . . . . 10-3 

10-2 Comparison of solar flux above the atmosphere with flux at the Earth’s surface . . . 10-8 

10-3 Ozone column abundances from the years 1990 to 1992 for 0, 40, and 80° N 

as well as 80° S . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 10-9 

10-4 Monthly averaged vertical O profiles (partial pressure in mPa) as a function 3

of atmospheric pressure for Trinidad Head, CA; Boulder, CO; Huntsville, AL 

and Wallops Island, VA .............................................. 10-11 

10-5 The sensitivity of ground-level UV flux to a 1 DU change in total column O , 3

under clear sky conditions, as a function of solar zenith angle (SZA) . . . . . . . . . . . 10-13 

10-6 Estimated global mean radiative forcing exerted by gas and various particle 

phase species for the year 2000, relative to 1750 . . . . . . . . . . . . . . . . . . . . . . . . . . . 10-47 

10-7 Mid-tropospheric O abundance (ppb) in northern midlatitudes (36 °N-59 °N) 3

for the years 1970 to 1996 ............................................ 10-52 

11-1 In-service fading of nylon 6 yarn inside house . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11-6 

11-2 In-service fading of nylon 6 yarn outside house . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11-7 

I-xxvi 
List of Figures 

(cont’d) 

Number	 Page 

11-3	 Observed color changes for natural colorant-on-paper systems during exposure 

to 0.40 ppm O at 25 °C ± 1 °C, 50% RH, in the absence of light 3 . . . . . . . . . . . . . . . 11-9

11-4	 Observed color changes for natural colorant-on-site during exposure to 

0.40 ppm O at 25 °C ± 1 °C, 50% RH, in the absence of light 3 . . . . . . . . . . . . . . . . 11-10

I-xxvii 
Authors, Contributors, and Reviewers 

CHAPTER 1. INTRODUCTION 

Principal Author 

Dr. Lester D. Grant—National Center for Environmental Assessment (B243-01), 

U.S. Environmental Protection Agency, Research Triangle Park, NC 27711 

CHAPTER 2 - PHYSICS AND CHEMISTRY OF OZONE IN THE ATMOSPHERE 

Principal Authors 

Dr. Joseph Pinto—National Center for Environmental Assessment (B243-01), 

U.S. Environmental Protection Agency, Research Triangle Park, NC 27711 

Dr. Russell Dickerson—Department of Atmospheric and Oceanic Sciences, University of 

Maryland, College Park, MD 

Contributing Authors 

Dr. Brooke Hemming—National Center for Environmental Assessment (B243-01), 

U.S. Environmental Protection Agency, Research Triangle Park, NC 27711 

Dr. Daniel Jacob—Department of Earth and Planetary Sciences, Harvard University, 

Cambridge, MA 

Dr. William Keene—Department of Environmental Sciences, University of Virginia, 

Charlottesville, VA 

Dr. Tadeusz Kleindienst—National Exposure Research Laboratory, U.S. Environmental 

Protection Agency, Research Triangle Park, NC 

Dr. Jennie Moody—Department of Environmental Sciences, University of Virginia, 

Charlottesville, VA 

Mr. Charles Piety—Department of Atmospheric and Oceanic Sciences, University of Maryland, 

College Park, MD 

Dr. Sandy Sillman—Department of Atmospheric, Oceanic, and Space Sciences, University of 

Michigan, Ann Arbor, MI 

I-xxviii 
Authors, Contributors, and Reviewers 

(cont’d) 

Contributing Authors 

(cont’d) 

Dr. Jeffrey Stehr—Department of Atmospheric and Oceanic Sciences, University of Maryland, 

College Park, MD 

Dr. Bret Taubman—Department of Atmospheric Sciences, Pennsylvania State University, 

State College, PA 

Contributors and Reviewers 

Dr. Christoph Bruhl—Max Planck Institute for Atmospheric Chemistry, Mainz, Germany 

Dr. Mohammed Elshahawy—Department of Meteorology and Astronomy, Cairo University,


Giza, Egypt.


Dr. Arlene Fiore—National Oceanic and Atmospheric Administration/Geophysical Fluid Dynamics


Laboratory, Princeton, NJ 


Mr. Chris Geron—National Risk Management Research Laboratory, U.S. Environmental Protection


Agency, Research Triangle Park, NC


Dr. David Golden—Department of Chemistry, Stanford University, Palo Alto, CA


Dr. John Merrill—Graduate School of Oceanography, University of Rhode Island, Kingston, RI


Dr. Sam Oltmans—National Oceanic and Atmospheric Administration/Climate Monitoring and


Diagnostic Laboratory, Boulder, CO


Dr. David Parrish—National Oceanic and Atmospheric Administration/Aeronomy Laboratory,


Boulder, CO


Dr. Perry Samson—Department of Atmospheric, Ocean, and Space Sciences, University of


Michigan, Ann Arbor, MI


Dr. Sandy Sillman—Department of Atmospheric, Ocean, and Space Sciences, University of


Michigan, Ann Arbor, MI


Dr. Melvin Shapiro—National Center for Atmospheric Research, Boulder, CO


I-xxix 
Authors, Contributors, and Reviewers 

(cont’d) 

CHAPTER 3 - ENVIRONMENTAL CONCENTRATIONS, PATTERNS, 

AND EXPOSURE ESTIMATES 

Principal Authors 

Dr. Joseph Pinto—National Center for Environmental Assessment (B243-01), 

U.S. Environmental Protection Agency, Research Triangle Park, NC 27711 

Ms. Beverly Comfort—National Center for Environmental Assessment (B243-01), 

U.S. Environmental Protection Agency, Research Triangle Park, NC 27711 

Dr. Arlene Fiore—National Oceanic and Atmospheric Administration/Geophysical Fluid Dynamics 

Laboratory, Princeton, NJ 

Dr. Daniel Jacob—Department of Earth and Planetary Sciences, Harvard University, 

Cambridge, MA 

Dr. Alan S. Lefohn—ASL &Associates, Helena, MT 

Dr. Clifford Weisel—Environmental and Occupational Health Sciences Institute, Rutgers 

University, New Brunswick, NJ 

Contributing Authors 

Dr. Jee Young Kim——National Center for Environmental Assessment (B243-01), 

U.S. Environmental Protection Agency, Research Triangle Park, NC 27711 

Dr. Dennis Kotchmar—National Center for Environmental Assessment (B243-01), 

U.S. Environmental Protection Agency, Research Triangle Park, NC 27711 

Dr. Timothy Lewis—National Center for Environmental Assessment (B243-01), 

U.S. Environmental Protection Agency, Research Triangle Park, NC 27711 

Mr. Thomas McCurdy—National Exposure Research Laboratory, U.S. Environmental Protection 

Agency, Research Triangle Park, NC 

Contributors and Reviewers 

Dr. Christoph Bruehl—Max Planck Institute for Atmospheric Chemistry, Mainz, Germany 

Dr. Russell Dickerson—Department of Atmospheric and Oceanic Sciences, University of 

Maryland, College Park, MD 

I-xxx 
Authors, Contributors, and Reviewers 

(cont’d) 

Contributors and Reviewers 

(cont’d) 

Dr. Judith Graham—American Chemistry Council, Washington, D.C. 

Dr. Laszlo Horvath—Hungarian Meteorological Service, Budapest, Hungary 

Dr. Ted Johnson—TRJ Associates, Durham, NC 

Dr. John Merrill—Graduate School of Oceanography, University of Rhode Island, Kingston, RI 

Dr. Jennie Moody—Department of Environmental Sciences, University of Virginia, 

Charlottesville, VA 

Dr. Sam Oltmans—National Oceanic and Atmospheric Administration/Climate Monitoring and 

Diagnostic Laboratory, Boulder, CO 

Dr. Michiel G.M. Roemer, TNO, The Netherlands 

Dr. Sandy Sillman—Department of Atmospheric, Ocean, and Space Sciences, University of 

Michigan, Ann Arbor, MI 

Dr. Tamas Weidinger—Department of Meteorology, University of Budapest, Budapest, Hungary 

CHAPTER 4 - DOSIMETRY, SPECIES HOMOLOGY, SENSITIVITY, 

AND EXTRAPOLATION 

Principal Authors 

Dr. John Overton—U.S. Environmental Protection Agency, National Health and Environmental 

Effects Research Laboratory-Research Triangle Park, NC 27711 (retired) 

Dr. James S. Brown—National Center for Environmental Assessment (B243-01), 

U.S. Environmental Protection Agency, Research Triangle Park, NC 27711 

Dr. Lori White—National Center for Environmental Assessment (B243-01), 

U.S. Environmental Protection Agency, Research Triangle Park, NC 27711 

I-xxxi 
Authors, Contributors, and Reviewers 

(cont’d) 

Contributors and Reviewers 

Dr. Gary Hatch—U.S. Environmental Protection Agency, National Health and Environmental 

Effects Research Laboratory, NC 

CHAPTER 5 - TOXICOLOGICAL EFFECTS IN LABORATORY ANIMALS AND 

IN VITRO TEST SYSTEMS 

Principal Authors 

Dr. Lori White—National Center for Environmental Assessment (B243-01), U.S. Environmental 

Protection Agency, Research Triangle Park, NC 27711 

Mr. James Raub—National Center for Environmental Assessment (B243-01), 

U.S. Environmental Protection Agency, Research Triangle Park, NC 27711 (retired) 

Dr. Deepak Bhalla—Department of Occupational and Environmental Health Sciences, Wayne State 

University, Detroit, MI 

Dr. Carroll Cross—School of Medicine, University of California, Davis, CA 

Dr. Mitch Cohen—NYU School of Medicine, New York University, New York, NY 

Contributors and Reviewers 

Dr. Steven Kleeberger—National Institute of Environmental Health Sciences, Research 

Triangle Park, NC 27711 

Dr. George Liekauf—Department of Environmental Health, University of Cincinnati, 

Cincinnati, OH 

Dr. David Basset—Department of Occupational and Environmental Health Sciences, Wayne State 

University, Detroit, MI 

Dr. E.M. Postlethwait—Department of Environmental Health Sciences, University of Texas 

Medical Branch, Galveston, TX 

Dr. Kent Pinkerton—Center for Health and the Environment, University of California, Davis, CA 

I-xxxii 
Authors, Contributors, and Reviewers 

(cont’d) 

Contributors and Reviewers 

(cont’d) 

Dr. Edward Schelegle—Department of Anatomy, Physiology, and Cell Biology, University of 

California, Davis, CA 

Dr. Judith Graham—American Chemical Council, Arlington, VA 

Dr. Paul Reinhart—National Center for Environmental Assessment (B243-03), U.S. Environmental 

Protection Agency, Research Triangle Park, NC 27711 

CHAPTER 6 - CONTROLLED HUMAN EXPOSURE STUDIES 

Principal Authors 

Dr. James S. Brown—National Center for Environmental Assessment (B243-01), 

U.S. Environmental Protection Agency, Research Triangle Park, NC 27711 

Mr. James Raub—National Center for Environmental Assessment (B243-01), U.S. Environmental 

Protection Agency, Research Triangle Park, NC 27711 (retired) 

Dr. William C. Adams—Human Performance Laboratory, University of California, Davis, CA 

(retired) 

Dr. Milan J. Hazucha—Center for Environmental Medicine, Asthma, and Lung Biology, 

University of North Carolina, Chapel Hill, NC 

Dr. E. William Spannhake—Department of Environmental Health Sciences, Johns Hopkins 

University, Baltimore, MD 

Contributors and Reviewers 

Dr. Edward Avol—Department of Preventive Medicine, University of Southern California, 

Los Angeles, CA 

Dr. Jane Q. Koenig—Department of Environmental and Occupational Health, University of 

Washington, Seattle, WA 

Dr. Michael Madden—National Health and Environmental Effects Research Laboratory, 

U.S. Environmental Protection Agency, Chapel Hill, NC 

I-xxxiii 
Authors, Contributors, and Reviewers 

(cont’d) 

Contributors and Reviewers 

(cont’d)


Dr.William McDonnell—National Health and Environmental Effects Research Laboratory,


U.S. Environmental Protection Agency, Chapel Hill, NC 

CHAPTER 7 - EPIDEMIOLOGICAL STUDIES OF HUMAN HEALTH EFFECTS 

Principal Authors 

Dr. Dennis Kotchmar—National Center for Environmental Assessment (B243-01), 

U.S. Environmental Protection Agency, Research Triangle Park, NC 27711 

Dr. Jee Young Kim—National Center for Environmental Assessment (B243-01), 

U.S. Environmental Protection Agency, Research Triangle Park, NC 27711 

Dr. David Svendsgaard—National Center for Environmental Assessment (B243-01), 

U.S. Environmental Protection Agency, Research Triangle Park, NC 27711 

Dr. Kazuhiko Ito—New York University School of Medicine, Nelson Institute of 

Environmental Medicine, Tuxedo, NY 

Dr. Patrick Kinney—Columbia University, Mailman School of Public Health, New York, NY 

Reviewers 

Dr. Richard Burnett—Health Canada, Ottawa, CN 

Dr. Vic Hasselblad—Duke University, Durham, NC 

Dr. Lucas Neas—National Health and Environmental Effects Research Laboratory, 

U.S. Environmental Protection Agency, Chapel Hill, NC 

I-xxxiv 
Authors, Contributors, and Reviewers 

(cont’d) 

CHAPTER 8 - INTEGRATIVE SYNTHESIS: EXPOSURE AND HEALTH EFFECTS 

Principal Authors 

Dr. Srikanth Nadadur—National Center for Environmental Assessment (B243-01), 

U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. Lester Grant—National Center for Environmental Assessment (B243-01), U.S. Environmental

Protection Agency, Research Triangle Park, NC 27711

Contributing Authors 

Dr. Jee Young Kim—National Center for Environmental Assessment (B243-01), 

U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. Joseph Pinto—National Center for Environmental Assessment (B243-01), 

U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. Mary Ross—National Center for Environmental Assessment (B243-01), U.S. Environmental

Protection Agency, Research Triangle Park, NC 27711

Dr. Lori White—National Center for Environmental Assessment (B243-01), U.S. Environmental

Protection Agency, Research Triangle Park, NC 27711

Dr. James S. Brown—National Center for Environmental Assessment (B243-01),

U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Reviewers 

Dr. John Vandenberg—National Center for Environmental Assessment, Washington, DC 

Dr. Daniel Costa—National Program Director for Air, Office of Research and Development,

Research Triangle Park, NC 27711

Dr. Paul Reinhart—National Center for Environmental Assessment (B243-03), U.S. Environmental

Protection Agency, Research Triangle Park, NC 27711

I-xxxv 
Authors, Contributors, and Reviewers 

(cont’d) 

CHAPTER 9 - ENVIRONMENTAL EFFECTS ON VEGETATION AND ECOSYSTEMS 

Principal Authors 

Dr. Jay Garner—National Center for Environmental Assessment (B243-01), U.S. Environmental 

Protection Agency, Research Triangle Park, NC 27711 (retired) 

Dr. Timothy Lewis—National Center for Environmental Assessment (B243-01), 

U.S. Environmental Protection Agency, Research Triangle Park, NC 27711 

Dr. William Hogsett—National Health and Environmental Effects Research Laboratory, 

U.S. Environmental Protection Agency, Corvallis, OR 

Dr. Christian Andersen—National Health and Environmental Effects Research Laboratory, 

U.S. Environmental Protection Agency, Corvallis, OR 

Dr. Allen Lefohn—ASL and Associates, Helena, MT 

Dr. David Karnosky—Forest Resources and Environmental Sciences, Michigan Technological 

University, Houghton, MI 

Dr. Michael Nannini—Center for Aquatic Ecology, Illinois Natural History Survey, Kinmundy, IL 

Dr. Nancy Grulke—Pacific Southwest Research Station Forest Fire Laboratory, USDA Forest 

Service, Riverside, CA 

Dr. Richard Adams—Department of Agriculture and Resource Economics, Oregon State 

University., Corvallis, OR 

Dr. Robert Heath—Department of Botany and Plant Sciences, University of California, 

Riverside, CA 

Dr. Victor Runeckle—Biology Department, University of British Columbia, Vancouver, B.C., CN 

(retired) 

Dr. Arthur Chappelka—Auburn University, School of Forestry, Auburn, AL 

Dr. William Massman—USDA Forest Service, Ft. Collins, CO 

Dr. Robert Musselman—USDA Forest Service, Fort Collins, CO 

Dr. Peter Woodbury—Cornell University, Ithaca, NY (former USDA Forest Service) 

I-xxxvi 
27603 

Authors, Contributors, and Reviewers 

(cont’d) 

Contributors and Reviewers 

Dr. Fitzgerald Booker—USDA-ARS Plant Science Research Unit, 3908 Inwood Rd., Raleigh, NC 

Dr. Boris Chevone—Department of Plant Pathology, Virginia Technological University, 

Blacksburg, VA 24061


Dr. Alan Davison—School of Biology, Newcastle University, Newcastle on Tyne,


United Kingdom, NE1 7RU


Dr. Bruce L. Dixon—Department of Agricultural Economics, University of Arkansas,


Fayetteville, AR 72701


Dr. David Grantz—Kearney Agricultural Center, University of California at Riverside, Parlier,


CA 93648


Dr. Allen S. Heagle—1216 Scott Pl., Raleigh, NC 27511 

Dr. Robert Horst, Jr.—121 Thorwald Dr., Plainsboro, NJ 08536 

Dr. John Innes—Forest Sciences Centre, Department of Forest Resources, University of British 

Columbia, Vancouver, BC, Canada V6T 1Z4 

Dr. Hans-Jürgen Jäger—Heinrich-Buff-Ring 26-32, Institute of Plant Ecology, Justus-Leibig


University, Gessen, Germany D35392


Dr. Robert Kohut— Tower Road, Boyce Thompson Institute, Rm 131,Cornell University,


Ithaca, NY 14853


Dr. Sagar Krupa—1519 Gortner Ave., Department of Plant Pathology, University of Minnesota,


St. Paul, MN 55108


Dr. William Manning—203 Morrill, Department of Microbiology, University of Massachusetts,


Amherst, MA 01003 

Dr. Howard Neufeld—Rankin Science Bldg., Appalachian State University, Boone, NC 28608 

Dr. Paul Miller—USDA Forest Service, Pacific Southwest Research Station, 4955 Canyon Crest 

Drive, Riverside CA 92507 

Dr. Maria-Jose Sanz—Fundacion CEAM, c/Charles Darein, 14-Parque Te, Valencia, Spain 

I-xxxvii 
Authors, Contributors, and Reviewers 

(cont’d) 

Contributors and Reviewers 

(cont’d) 

Dr. James Shortle—Department of Ag Econ, Armsby, Pennsylvania State University, 

University Park, PA 16802 

Dr. John Skelly—Department of Plant Pathology, Pennsylvania State University, University 

Park, PA 16803 

CHAPTER 10 - TROPOSPHERIC OZONE EFFECTS ON UV-B FLUX 

AND CLIMATE CHANGE 

Principal Authors 

Dr. Brooke Hemming—National Center for Environmental Assessment (B243-01), 

U.S. Environmental Protection Agency, Research Triangle Park, NC 27711 

Dr. Jee Young Kim—National Center for Environmental Assessment (B243-01), 

U.S. Environmental Protection Agency, Research Triangle Park, NC 27711 

Contributors and Reviewers 

Dr. Sasha Madronich—Atmospheric Chemistry Division. National Center for Atmospheric 

Research (NCAR), Boulder, CO 80307 

Dr. Daniel J. Jacob—Atmospheric Chemistry and Environmental Engineering, Division of 

Engineering & Applied Science, and Department of Earth & Planetary Sciences, Harvard 

University, Cambridge, MA 02138 

CHAPTER 11 - EFFECTS OF OZONE ON MAN-MADE MATERIALS 

Principal Author 

Mr. Bill Ewald—National Center for Environmental Assessment (B243-01), U.S. Environmental 

Protection Agency, Research Triangle Park, NC 27711 (retired) 

I-xxxviii 
U.S. Environmental Protection Agency Project Team 

for Development of Air Quality Criteria for Ozone 

and Related Photochemical Oxidants 

Executive Direction 

Dr. Lester D. Grant (Director)—National Center for Environmental Assessment-RTP Division,


(B243-01), U.S. Environmental Protection Agency, Research Triangle Park, NC 27711


Scientific Staff 

Dr. Lori White(Ozone Team Leader)—National Center for Environmental Assessment (B243-01), 

U.S. Environmental Protection Agency, Research Triangle Park, NC 27711


Dr. Joseph Pinto—National Center for Environmental Assessment (B243-01), 


U.S. Environmental Protection Agency, Research Triangle Park, NC 27711


Ms. Beverly Comfort—National Center for Environmental Assessment (B243-01), 


U.S. Environmental Protection Agency, Research Triangle Park, NC 27711


Dr. Brooke Hemming—National Center for Environmental Assessment (B243-01),


U.S. Environmental Protection Agency, Research Triangle Park, NC 27711


Dr. James S. Brown—National Center for Environmental Assessment (B243-01), 


U.S. Environmental Protection Agency, Research Triangle Park, NC 27711


Dr. Dennis Kotchmar—National Center for Environmental Assessment (B243-01),


U.S. Environmental Protection Agency, Research Triangle Park, NC 27711


Dr. Jee Young Kim—National Center for Environmental Assessment (B243-01),


U.S. Environmental Protection Agency, Research Triangle Park, NC 27711


Dr. David Svendsgaard—National Center for Environmental Assessment (B243-01),


U.S. Environmental Protection Agency, Research Triangle Park, NC 27711


Dr. Srikanth Nadadur—National Center for Environmental Assessment (B243-01),


U.S. Environmental Protection Agency, Research Triangle Park, NC 27711


Dr. Timothy Lewis—National Center for Environmental Assessment (B243-01),


U.S. Environmental Protection Agency, Research Triangle Park, NC 27711 


Dr. Jay Garner—National Center for Environmental Assessment (B243-01), 


U.S. Environmental Protection Agency, Research Triangle Park, NC 27711 (retired) 

I-xxxix 
U.S. Environmental Protection Agency Project Team 

for Development of Air Quality Criteria for Ozone 

and Related Photochemical Oxidants 

(cont’d) 

Scientific Staff 

(cont’d) 

Dr. William Hogsett—National Health and Environmental Effects Research Laboratory, 

U.S. Environmental Protection Agency, Corvallis, OR 

Dr. Christian Andersen—National Health and Environmental Effects Research Laboratory, 

U.S. Environmental Protection Agency, Corvallis, OR 

Mr. Bill Ewald—National Center for Environmental Assessment (B243-01), 

U.S. Environmental Protection Agency, Research Triangle Park, NC 27711 (retired) 

Mr. James Raub—National Center for Environmental Assessment (B243-01), 

U.S. Environmental Protection Agency, Research Triangle Park, NC 27711 (retired) 

Technical Support Staff 

Ms. Nancy Broom—Information Technology Manager, National Center for Environmental 

Assessment (B243-01), U.S. Environmental Protection Agency, Research Triangle Park, NC 27711 

Mr. Douglas B. Fennell—Technical Information Specialist, National Center for Environmental 

Assessment (B243-01), U.S. Environmental Protection Agency, Research Triangle Park, NC 27711 

Ms. Emily R. Lee—Management Analyst, National Center for Environmental Assessment 

(B243-01), U.S. Environmental Protection Agency, Research Triangle Park, NC 27711 

Ms. Diane H. Ray—Program Specialist, National Center for Environmental Assessment 

(B243-01), U.S. Environmental Protection Agency, Research Triangle Park, NC 27711 

Ms. Donna Wicker—Administrative Officer, National Center for Environmental Assessment 

(B243-01), U.S. Environmental Protection Agency, Research Triangle Park, NC 27711 (retired) 

Mr. Richard Wilson—Clerk, National Center for Environmental Assessment (B243-01), 

U.S. Environmental Protection Agency, Research Triangle Park, NC 27711 

I-xl 
U.S. Environmental Protection Agency Project Team 

for Development of Air Quality Criteria for Ozone 

and Related Photochemical Oxidants 

(cont’d) 

Document Production Staff 

Ms. Carolyn T. Perry—Manager, Computer Sciences Corporation, 2803 Slater Road, Suite 220, 

Morrisville, NC 27560 

Mr. John A. Bennett—Technical Information Specialist, Library Associates of Maryland, 

11820 Parklawn Drive, Suite 400, Rockville, MD 20852 

Mr. William Ellis—Records Management Technician, InfoPro, Inc., 8200 Greensboro Drive, 

Suite 1450, McLean, VA 22102 

Ms. Sandra L. Hughey—Technical Information Specialist, Library Associates of Maryland, 

11820 Parklawn Drive, Suite 400, Rockville, MD 20852 

Mr. Matthew Kirk—Graphic Artist, Computer Sciences Corporation, 2803 Slater Road, Suite 220, 

Morrisville, NC 27560 

Dr. Barbara Liljequist—Technical Editor, Computer Sciences Corporation, 2803 Slater Road, 

Suite 220, Morrisville, NC 27560 

Ms. Rosemary Procko—Senior Word Processor, TekSystems, 1201 Edwards Mill Road, Suite 201, 

Raleigh, NC 27607 

Ms. Faye Silliman—Publication/Graphics Specialist, InfoPro, Inc., 8200 Greensboro Drive, 

Suite 1450, McLean, VA 22102 

Mr. Carlton Witherspoon—Graphic Artist, Computer Sciences Corporation, 2803 Slater Road, 

Suite 220, Morrisville, NC 27560 

I-xli 
U.S. Environmental Protection Agency Science Advisory Board (SAB) 

Staff Office Clean Air Scientific Advisory Committee (CASAC) 

Ozone Review Panel 

Chair 

Dr. Rogene Henderson*, Scientist Emeritus, Lovelace Respiratory Research Institute, 2425 

Ridgecrest Drive SE, Albuquerque, NM, 87108, Phone: 505-348-9464, Fax: 505-348-8541, 

(rhenders@lrri.org) (FedEx: Dr. Rogene Henderson, Lovelace Respiratory Research Institute, 

2425 Ridgecrest Drive SE, Albuquerque, NM, 87108, Phone: 505-348-9464) 

Members 

Dr. John Balmes, Professor, Department of Medicine, University of California San Francisco, 

University of California - San Francisco, San Francisco, California, 94143, Phone: 415-206-8953, 

Fax: 415-206-8949, (jbalmes@itsa.ucsf.edu) 

Dr. Ellis Cowling*, University Distinguished Professor-at-Large, North Carolina State University, 

Colleges of Natural Resources and Agriculture and Life Sciences, North Carolina State University, 

1509 Varsity Drive, Raleigh, NC, 27695-7632, Phone: 919-515-7564 , Fax: 919-515-1700, 

(ellis_cowling@ncsu.edu) 

Dr. James D. Crapo*, Professor, Department of Medicine, National Jewish Medical and Research 

Center. 1400 Jackson Street, Denver, CO, 80206, Phone: 303-398-1436, Fax: 303- 270-2243, 

(crapoj@njc.org) 

Dr. William (Jim) Gauderman, Associate Professor, Preventive Medicine, University of 

Southerm California, 1540 Alcazar #220, Los Angeles, CA, 91016, Phone: 323-442-1567, 

Fax: 323-442-2349, (jimg@usc.edu) 

Dr. Henry Gong, Professor of Medicine and Preventive Medicine, Medicine and Preventive 

Medicine, Keck School of Medicine, University of Southern California, Environmental Health 

Service, MSB 51, Rancho Los Amigos NRC, 7601 East Imperial Highway, Downey, CA, 90242, 

Phone: 562-401-7561, Fax: 562-803-6883, (hgong@ladhs.org) 

Dr. Paul J. Hanson, Senior Research and Development Scientist , Environmental Sciences Division, 

Oak Ridge National Laboratory (ORNL), Bethel Valley Road, Building 1062, Oak Ridge, TN, 

37831-6422, Phone: 865-574-5361, Fax: 865-576-9939, (hansonpz@comcast.net) 

Dr. Jack Harkema, Professor, Department of Pathobiology, College of Veterinary Medicine, 

Michigan State University, 212 Food Safety & Toxicology Center, East Lansing, MI, 48824, 

Phone: 517-353-8627, Fax: 517-353-9902, (harkemaj@msu.edu) 

I-xlii 
U.S. Environmental Protection Agency Science Advisory Board (SAB) 

Staff Office Clean Air Scientific Advisory Committee (CASAC) 

Ozone Review Panel 

(cont’d) 

Members 

(cont’d) 

Dr. Philip Hopke, Bayard D. Clarkson Distinguished Professor, Department of Chemical 

Engineering, Clarkson University, Box 5708, Potsdam, NY, 13699-5708, Phone: 315-268-3861, 

Fax: 315-268-4410, (hopkepk@clarkson.edu) (FedEx: 8 Clarkson Avenue, Potsdam, NY 

136995708) 

Dr. Michael T. Kleinman, Professor, Department of Community & Environmental Medicine, 

100 FRF, University of California - Irvine, Irvine, CA, 92697-1825, Phone: 949-824-4765, Fax: 

949-824-2070, (mtkleinm@uci.edu) 

Dr. Allan Legge, President, Biosphere Solutions, 1601 11th Avenue NW, Calgary, Alberta, 

CANADA, T2N 1H1, Phone: 403-282-4479, Fax: 403-282-4479, (allan.legge@shaw.ca) 

Dr. Morton Lippmann, Professor, Nelson Institute of Environmental Medicine, New York 

University School of Medicine, 57 Old Forge Road, Tuxedo, NY, 10987, Phone: 845-731-3558, 

Fax: 845-351-5472, (lippmann@env.med.nyu.edu) 

Dr. Frederick J. Miller*, Consultant, 911 Queensferry Road, Cary, NC, 27511, Phone: 

919-467-3194, (fjmiller@nc.rr.com) 

Dr. Maria Morandi, Assistant Professor of Environmental Science & Occupational Health, 

Department of Environmental Sciences, School of Public Health, University of Texas - Houston 

Health Science Center, 1200 Herman Pressler Street, Houston, TX, 77030, Phone: 713-500-9288, 

Fax: 713-500-9249, (mmorandi@sph.uth.tmc.edu) (FedEx: 1200 Herman Pressler, Suite 624) 

Dr. Charles Plopper, Professor, Department of Anatomy, Physiology and Cell Biology, School of 

Veterinary Medicine, University of California - Davis, Davis, California, 95616, Phone: 

530-752-7065, (cgplopper@ucdavis.edu) 

Mr. Richard L. Poirot*, Environmental Analyst, Air Pollution Control Division, Department of 

Environmental Conservation, Vermont Agency of Natural Resources, Bldg. 3 South, 103 South 

Main Street, Waterbury, VT, 05671-0402, Phone: 802-241-3807, Fax: 802-241-2590, 

(rich.poirot@state.vt.us) 

Dr. Armistead (Ted) Russell, Georgia Power Distinguished Professor of Environmental 

Engineering, Environmental Engineering Group, School of Civil and Environmental Engineering, 

Georgia Institute of Technology, 311 Ferst Drive, Room 3310, Atlanta, GA, 30332-0512, Phone: 

404-894-3079, Fax: 404-894-8266, (trussell@ce.gatech.edu) 

I-xliii 
U.S. Environmental Protection Agency Science Advisory Board (SAB) 

Staff Office Clean Air Scientific Advisory Committee (CASAC) 

Ozone Review Panel 

(cont’d) 

Members 

(cont’d) 

Dr. Elizabeth A. (Lianne) Sheppard, Research Associate Professor, Biostatistics and Environmental 

& Occupational Health Sciences, Public Health and Community Medicine, University of 

Washington, Box 357232, Seattle, WA, 98195-7232, Phone: 206-616-2722, Fax: 206 616-2724, 

(sheppard@u.washington.edu) 

Dr. Frank Speizer*, Edward Kass Professor of Medicine, Channing Laboratory, Harvard Medical 

School, 181 Longwood Avenue, Boston, MA, 02115-5804, Phone: 617-525-2275, Fax: 

617-525-2066, (frank.speizer@channing.harvard.edu) 

Dr. James Ultman, Professor, Chemical Engineering, Bioengineering program, Pennsylvania State 

University, 106 Fenske Lab, University Park, PA, 16802, Phone: 814-863-4802, Fax: 

814-865-7846, (jsu@psu.edu) 

Dr. Sverre Vedal, Professor of Medicine, Department of Environmental and Occupational Health 

Sciences, School of Public Health and Community Medicine, University of Washington, 4225 

Roosevelt Way NE, Suite 100, Seattle, WA, 98105-6099, Phone: 206-616-8285, Fax: 

206-685-4696, (svedal@u.washington.edu) 

Dr. James (Jim) Zidek, Professor, Statistics, Science, University of British Columbia, 6856 

Agriculture Rd., Vancouver, BC, Canada, V6T 1Z2, Phone: 604-822-4302, Fax: 604-822-6960, 

(jim@stat.ubc.ca) 

Dr. Barbara Zielinska*, Research Professor , Division of Atmospheric Science, Desert Research 

Institute, 2215 Raggio Parkway, Reno, NV, 89512-1095, Phone: 775-674-7066, Fax: 

775-674-7008, (barbz@dri.edu) 

Science Advisory Board Staff 

Mr. Fred Butterfield, CASAC Designated Federal Officer, 1200 Pennsylvania Avenue, N.W., 

Washington, DC, 20460, Phone: 202-343-9994, Fax: 202-233-0643 (butterfield.fred@epa.gov) 

(Physical/Courier/FedEx Address: Fred A. Butterfield, III, EPA Science Advisory Board Staff 

Office (Mail Code 1400F), Woodies Building, 1025 F Street, N.W., Room 3604, Washington, 

DC 20004, Telephone: 202-343-9994) 

*Members of the statutory Clean Air Scientific Advisory Committee (CASAC) appointed by

 the EPA Administrator 

I-xliv 
A/V 

ABBREVIATIONS AND ACRONYMS 

" alpha; probability value 

AA ascorbic acid 

ACh acetylcholine 

ADSS aged and diluted cigarette smoke 

AER air exchange rate 

AEROCE Atmospheric/Ocean Chemistry Experiment 

AHR airway hyperreactivity 

AHSMOG Adventist Health Study on Smog 

AIRPEX Air Pollution Exposure (model) 

AIRQUIS Air Quality Information System (model) 

AIRS Aerometric Information Retrieval System 

AM alveolar macrophage 

ANF atrial natriuretic factor 

AOP2 antioxidant protein 2 

AOT40 seasonal sum of the difference between an hourly concentration at the 

threshold value of 40 ppb, minus the threshold value of 40 ppb 

APEX	 Air Pollution Exposure (model) 

APHEA	 Air Pollution on Health: European Approach (study) 

AQCD	 Air Quality Criteria Document 

AQS	 Air Quality System 

ARIC	 Atherosclerosis Risk in Communities (study) 

ATS	 American Thoracic Society 

surface-to-volume ratio 

$ beta-coefficient; slope of an equation 

BAL	 bronchioalveolar lavage 

BALF	 bronchioalveolar lavage fluid 

BC	 black carbon 

BLD	 below limit of detection 

I-xlv 
BS black smoke 

BSA body surface area 

BSA bovine serum albumin 

BMZ basement membrane zone 

BP blood pressure 

C concentration 

C3a complement protein fragment 

CAA Clean Air Act 

CADS Cincinnati Activity Diary Study 

CAPs concentrated ambient particles 

CAR centriacinar region 

CASAC Clean Air Scientific Advisory Committee 

CASTNet, CASTNET Clean Air Status and Trends Network 

CC16 Clara cell secretory protein 

CCSP Clara cell secretory protein 

Cdyn , Cdyn dynamic lung compliance 

CDT Central Daylight Time 

CE continuous exercise 

CFCs chlorofluorocarbons 

CFD computational fluid dynamics 

CFR Code of Federal Regulations 

CH4 methane 

C H –H 2 5 ethane 

C H 5 8 isoprene 

C H 10 16 terpene 

CHAD Consolidated Human Activities Database 

CH –CCl 3 3 methyl chloroform 

CH –CHO 3 acetaldehyde 

CH –CO 3 acetyl 

CHO Chinese hamster ovary (cells) 

I-xlvi 
CH OOH 3

CI 

CIE 

CINC 

CLM 

CMAQ 

CO 

CO2

COD 

COP 

COPD 

CRP 

CTM 

)

3-D 

DHBA 

DNA 

DOAS 

DPPC 

DTPA 

DU 

,

ECG 

EDU 

EEG 

ELF 

ENA-78 

ENSO 

EPA 

EST 

acetic acid 

confidence interval 

Commission Internationale de l’Eclaiarage (International Commission on 

Illumination) 

cytokine-induced neutrophil chemoattractant 

chemiluminescence method 

Community Model for Air Quality 

carbon monoxide 

carbon dioxide 

coefficient of divergence 

Conference of Parties 

chronic obstructive pulmonary disease 

C-reactive protein 

chemistry transport model 

delta; change in a variable 

three-dimensional 

2,3-dihydroxybenzoic acid 

deoxyribonucleic acid 

differential optical absorption spectroscopy 

dipalmitoylglycero-3-phosphocholine 

diethylenetriaminepentaacetic acid 

Dobson units 

epsilon; convergence precision 

electrocardiographic; electrocardiogram 

ethylenediurea 

electroencephalographic 

epithelial lining fluid 

epithelial cell-derived neutrophil-activating peptide 78 

El Niño-Southern Oscillation 

U.S. Environmental Protection Agency 

Eastern Standard Time 

I-xlvii 
ETS environmental tobacco smoke 

EVR equivalent ventilation rate 

F female 

F344 Fisher 344 (rat) 

FA filtered air 

FACE free-air carbon dioxide exposure 

fB breathing frequency 

FEF forced expiratory flow 

FEF25-75 forced expiratory flow between 25 and 75% of vital capacity; forced 

expiratory flow at 25 to 75% of vital capacity 

FEF x forced expiratory flow after X% vital capacity (e.g., after 25, 50, or 

75% vital capacity) 

FEV1 forced expiratory volume in 1 second 

FGF fibroblast growth factor 

FGFR fibroblast growth factor receptor 

FN fibronectin 

FR Federal Register 

FRM Federal Reference Method 

FVC forced vital capacity 

GAM Generalized Additive Model 

GCM general circulation model 

GEE Generalized Estimating Equation 

GEOS-CHEM three-dimensional model of atmospheric composition driven by 

assimilated Goddard Earth Orbiting System observations 

GHG greenhouse gas 

GLM Generalized Linear Model 

GLRAG Great Lakes Regional Assessment Group 

GM-CSF granulocyte-macrophage colony stimulating factor 

G6PD glucose-6-phosphate dehydrogenase 

GPx glutathione peroxidase 

GR glutathione reductase 

I-xlviii 
GSH glutathione; reduced glutathione 

GSHPx glutathione peroxidase 

GSTM1 glutathione S-transferase μ-1 (genotype) 

GSTM1null glutathione S-transferase μ-1 null (genotype) 

+ H hydrogen ion 

HCFCs hydrochlorofluorocarbons 

H CO, HCHO formaldehyde 2

HDMA house dust mite allergen 

HF hydrofluoride 

HFCs hydrofluorocarbons 

HLA human leukocyte antigen 

HNE 4-hydroxynonenal 

HNO , HONO nitrous acid 2 

HNO3 nitric acid 

HNO4 pernitric acid 

HO2 hydroperoxyl; hydroperoxy 

H O2 hydrogen peroxide 2 

HOx hydrogen oxides 

HR heart rate 

HRP horseradish peroxidase 

HRV heart rate variability 

H SO4 sulfuric acid 2 

IARC International Agency for Research on Cancer 

IC inspiratory capacity 

ICAM intracellular adhesion molecule 

ICNIRP International Commission on Non-Ionizing Radiation Protection 

IE intermittent exercise 

IFN interferon 

Ig immunoglobulin (e.g., IgA, IgE, IgG, IgM) 

IL interleukin (e.g., IL-1, ILIL-6, IL-8) 

I-xlix 
iNOS inducible nitric oxide synthase; NOS-2 

i.p. intraperitoneal 

IPCC Intergovernmental Panel on Climate Change 

IQR interquartile range 

IR infrared 

Ka intrinsic mass transfer coefficient/parameter 

K mass transfer coefficient for gas phase g

Kl mass transfer coefficient for liquid phase 

Kr reaction rate constant 

LDH lactic acid dehydrogenase 

LIDAR LIght Detection And Ranging 

LIS lateral intercellular space 

LLJ low-level jet 

LOEL lowest-observed-effect level 

LOESS locally estimated smoothing splines 

LOP lipid ozonzation products 

LPS lipopolysaccharide 

LRT lower respiratory tract; lower airways 

LT leukotriene (e.g., LTB , LTC , LTD , LTE4) 4 4 4 

LT local time 

LT" lymphotoxin-"

M male 

M maximum number of iterations 

MAP mean arterial pressure 

MARAT Mid-Atlantic Regional Assessment Team 

MCh methacholine 

MCP monocyte chemotactic protein 

MED minimal erythema dose 

MENTOR Modeling Environment for Total Risk Studies 

MET metabolic equivalent 

I-l 
MI myocardial infarction 

MIP macrophage inflammatory protein 

MMEF maximal midexpiratory flow 

MONICA Monitoring Trend and Determinants in Cardiovascular Disease 

(registry) 

MPAN peroxymethacryloyl nitrate; peroxy-methacrylic nitric anhydride 

MPO myeloperoxidase 

mRNA messenger ribonucleic acid 

MSA metropolitan statistical area 

MSL mean sea level 

MT metallothionein 

n, N number 

NAAQS National Ambient Air Quality Standards 

NAD(P)H reduced nicotinamide adenine dinucleotide phosphate 

NAMS/SLAMS National Ambient Monitoring Stations and State and Local Air 

Monitoring Stations 

NARSTO North American Regional Strategy for Atmospheric Ozone 

NAS Normative Aging Study 

NCEA-RTP National Center for Environmental Assessment Division in Research 

Triangle Park, NC 

NCICAS National Cooperative Inner-City Asthma Study 

NCLAN National Crop Loss Assessment Network 

ND not detectable; not detected 

NEM National Ambient Air Quality Standards Exposure Model 

NERAG New England Regional Assessment Group 

NF-6B nuclear factor kappa B 

NH3 ammonia 

NHAPS National Human Activity Pattern Survey 

NH HSO4 ammonium bisulfate 

(NH ) HSO4 ammonium sulfate 42 

NIH National Institutes of Health 

I-li 

4 
NIST National Institute of Standards and Technology 

NK natural killer (cells) 

NL nasal lavage 

NM national monument 

NMHCs nonmethane hydrocarbons 

NMMAPS National Morbidity, Mortality and Air Pollution Study 

NO nitric oxide 

NO2 nitrogen dioxide 

N O nitrous oxide 2

! NO3 nitrate 

N O5 nitrogen dioxide 2 

NOS nitric oxide synthase 

NOS-1 neuronal nitric oxide synthase 

NOS-2 inducible nitric oxide synthase; iNOS 

NOS-3 endothelial nitric oxide synthase 

NOx nitrogen oxides 

NO reactive nitrogen system components; sum of NO and NO ; odd y x z 

nitrogen species 

NOz difference between NO and NO ; reservoir and termination nitrogen y x 

species 

NP national park 

NPP net primary productivity 

NQO1 NAD(P)H-quinone oxidoreductase (genotype) 

NQO1wt NAD(P)H-quinone oxidoreductase wild type (genotype) 

NRC National Research Council 

NS national seashore 

NTP National Toxicology Program 

NTS nucleus tractus solitarius 

NWR national wildlife refuge 

O2 ground-state oxygen 

O3 ozone 

I-lii 
O * electronically excited ozone 3

18O3 radiolabeled ozone 

1 O( D) electronically excited oxygen atom 

OAQPS Office of Air Quality Planning and Standards 

OH hydroxyl; hydroxy 

8-OHdG 8-hydroxy-2N-deoxyguanosine 

3 

P 

O( P) ground-state oxygen atom 

OPE ozone production efficiency 

OTC open-top chamber 

OVA ovalbumin 

OxComp oxidative capacity of the atmosphere 

p probability value 

90 90th percentile 

PAF platelet-activating factor 

PAN peroxyacetyl nitrate; peroxyacetic nitric anhydride 

P O2 partial pressure of arterial oxygen a 

P 

PAR proximal alveolar region 

PBL planetary boundary layer 

PBPK physiologically based pharmacokinetic (approach) 

PCl picryl chloride 

PE postexposure 

PEF peak expiratory flow 

PEFR peak expiratory flow rate 

PEM personal exposure monitor 

enh enhanced pause 

PG prostaglandin (e.g., PGD , PGE, PGE , PGE PGF , PGF2") 2 1 2 1"

6PGD 6-phosphogluconate dehydrogenase 

PGP protein gene product (e.g., PGP9.5) 

PI probability interval 

PM particulate matter 

I-liii 
2 

PM 

PM2.5 fine particulate matter (mass median aerodynamic diameter #2.5 μm) 

PM10 combination of coarse and fine particulate matter 

10-2.5 coarse particulate matter (mass median aerodynamic diameter between 

10 and 2.5 μm) 

PMNs polymorphonuclear leukocytes; neutrophils 

pNEM Probabilistic National Ambient Air Quality Standard Exposure Model 

polyADPR poly(adenosinediphosphate-ribose) 

ppb parts per billion 

ppbv parts per billion by volume 

ppm parts per million 

PPN peroxypropionyl nitrate; peroxypropionic nitric anhydride 

PRB policy relevant background 

PSA picryl sulfonic acid 

PSC polar stratospheric clouds 

PUFA polyunsaturated fatty acid 

PWM pokeweed mitogen 

QCE quasi continuous exercise 

R intraclass correlation coefficient 

r correlation coefficient 

R multiple correlation coefficient 

Raw airway resistance 

R'CO acyl 

R'C(O)–O acyl peroxy 2 

RH relative humidity 

RL total pulmonary resistance 

RNA ribonucleic acid 

RO2 organic peroxyl; organic peroxy 

ROOH organic peroxides 

ROS reactive oxygen species 

RR ribonucleotide reductase 

RRMS relatively remote monitoring sites 

I-liv 
RT respiratory tract 

SAB Science Advisory Board 

SAC Staphylococcus aureus Cowan 1 strain 

SAMD S-adenosyl methionine decarboxylase 

S O2 oxygen saturation of arterial blood a 

SO 

SBUV solar backscattered ultraviolet radiation 

SC stratum corneum 

SD Sprague-Dawley (rat) 

SD standard deviation 

SES socioeconomic status 

sGAW specific airways conductance 

SHEDS Simulation of Human Exposure and Dose System 

sICAM soluble intracellular adhesion molecule 

SNAAQS Secondary National Ambient Air Quality Standards 

SNPs single nucleotide polymorphisms 

SO2 sulfur dioxide 

4

2! sulfate 

SOA secondary organic aerosol 

SOD superoxide dismutase 

SOS Southern Oxidant Study 

SP substance P 

SP surfactant protein (e.g., SP-A, SP-D) 

SPF specific-pathogen free 

sRAW, SRaw specific airways resistance 

SRES Special Report on Emissions Scenarios 

STE stratospheric-tropospheric exchange 

STRF Spatio-Temporal Random Field 

SUM06 seasonal sum of all hourly average concentrations $0.06 ppm 

SUM08 seasonal sum of all hourly average concentrations $0.08 ppm 

SZA solar zenith angle 

I-lv 
T time; duration of exposure 

t t-test statistical value; t statistic 

T3 triiodothyronine 

T4 thyroxine 

TAR Third Assessment Report 

TAR WGI Third Assessment Report of Working Group 1 

TB terminal bronchioles 

TBARS thiobarbituric acid reactive substances 

99mTc-DTPA radiolabeled diethylenetriaminepentaacetic acid 

TCO core temperature 

TCTL cytotoxic T-lymphocytes 

TiO2 titanium dioxide 

TLC total lung capacity 

TLR Toll-like receptor 

TNF tumor necrosis factor 

TNFR tumor necrosis factor receptor 

TOMS Total Ozone Mapping/Monitoring Satellite; total ozone mapping 

spectrometer 

TRIM Total Risk Integrated Methodology (model) 

TRIM.Expo Total Risk Integrated Methodology Exposure Event (model) 

TSP total suspended particulate 

TWA time-weighted average 

UA uric acid 

UNEP United Nations Environmental Program 

UNFCCC United Nations Framework Convention on Climate Change 

URT upper respiratory tract; upper airways 

USGCRP U.S. Global Change Research Program 

UV ultraviolet 

UV-A ultraviolet radiation of wavelengths 320 to 400 nm 

UV-B ultraviolet radiation of wavelengths 280 to 320 nm 

UV-C ultraviolet radiation of wavelengths 200 to 280 nm 

I-lvi 
VC vital capacity 

VD volume of the anatomic dead space 

VE minute ventilation; expired volume per minute 0

VO 0 2max maximal oxygen uptake (maximal aerobic capacity) 

VOC volatile organic compound 

VP volumetric penetration 0

V0 P50% volume at which 50% of an inhaled bolus is absorbed 

V 

VT tidal volume 

VTB terminal bronchiole region volume 

UA volume of the upper airways 

W126 cumulative integrated exposure index with a sigmoidal weighting function 

WMO World Meteorological Organization 

WT wild type 

I-lvii 
E-1

EXECUTIVE SUMMARY

E.1 INTRODUCTION

Tropospheric or “surface-level” ozone (O3) is one of six major air pollutants regulated by

National Ambient Air Quality Standards (NAAQS) under the U.S. Clean Air Act. As mandated

by the Clean Air Act, the U.S. Environmental Protection Agency (EPA) must periodically

review the scientific bases (or “criteria”) for the various NAAQS by assessing newly available

scientific information on a given criteria air pollutant. This document, Air Quality Criteria for

Ozone and Other Photochemical Oxidants, is an updated revision of the 1996 Ozone Air Quality

Criteria Document (O3 AQCD) that provided scientific bases for the current O3 NAAQS set in

1997. 

E.1.1 Clean Air Act Legal Requirements

Clean Air Act (CAA) Sections 108 and 109 govern establishment, review, and revision of

U.S. National Ambient Air Quality Standards (NAAQS). 

! Section 108 directs the U.S. Environmental Protection Agency (EPA) Administrator to list

ubiquitous (widespread) air pollutants that may reasonably be anticipated to endanger public

health or welfare and to issue air quality criteria for them. The air quality criteria are to

reflect the latest scientific information useful in indicating the kind and extent of all

exposure-related effects on public health and welfare expected from the presence of the

pollutant in the ambient air.

! Section 109 directs the EPA Administrator to set and periodically revise, as appropriate, two

types of NAAQS: (a) primary NAAQS to protect against adverse health effects of listed

criteria pollutants among sensitive population groups, with an adequate margin of safety, and

(b) secondary NAAQS to protect against welfare effects (e.g., impacts on vegetation, crops,

ecosystems, visibility, climate, man-made materials, etc.). Section 109 also requires peer

review of the NAAQS and their underlying scientific bases by the Clean Air Scientific

Advisory Committee (CASAC), a committee of independent non-EPA experts. 
E-2

E.1.2 Chronology of Ozone NAAQS Revisions

In 1971, the U.S. EPA set primary and secondary standards for total photochemical

oxidants. Based on a criteria review completed in 1978, the original NAAQS set in 1971 were

revised in 1979 to focus on O3 as the indicator for new primary and secondary standards that

would be attained when the expected number of days per calender year with maximum 1-h

average O3 concentrations >0.12 ppm did not exceed one. The NAAQS for ambient O3 were

revised in 1997 by replacing the 1-h standards with an 8-h primary standard that is met when the

3-year average of the annual fourth highest daily maximum 8-h average concentration is

#0.08 ppm. The 1997 primary NAAQS was based on scientific data from controlled human

exposure, laboratory animal, and epidemiological studies and associated analyses presented in

the 1996 O3 AQCD and in the 1996 O3 Staff Paper. 

! This revised O3 AQCD, prepared by EPA=s National Center for Environmental Assessment

(NCEA), provides scientific bases to support the periodic review of O3 NAAQS. This

document assesses the latest available scientific information (published mainly through

December 2004) judged to be useful in deriving criteria as scientific bases for decisions on

possible revision of the current O3 NAAQS.

! A separate EPA O3 Staff Paper, prepared by EPA’s Office of Air Quality Planning and

Standards (OAQPS), will draw upon key findings/conclusions from this document, together

with other analyses, to develop and present options for consideration by the EPA

Administrator regarding review and possible revision of the O3 NAAQS. 

E.1.3 Document Organization and Structure

Volume I of this document consists of the present Executive Summary and eleven main

chapters of this revised O3 AQCD. Those main chapters focus primarily on interpretative

evaluation of key information, whereas more detailed descriptive summarization of pertinent

studies and/or supporting analyses are provided in accompanying annexes. Volume II contains

the annexes for Chapters 4 through 7, whereas Volume III contains the annex for Chapter 9.
E-3

Topics covered in the main chapters of the present AQCD are as follows:

! This Executive Summary summarizes key findings and conclusions from Chapters 1 through

11 of this revised O3 AQCD, as they pertain to background information on O3-related

atmospheric science and air quality, human exposure aspects, dosimetric considerations,

health effect issues, and environmental effect issues. 

! Chapter 1 provides a general introduction, including an overview of legal requirements,

the chronology of past revisions of O3-related NAAQS, and orientation to the structure of

this document. 

! Chapters 2 and 3 provide background information on atmospheric chemistry/physics of O3

formation, air quality, and exposure aspects to help to place ensuing discussions of O3 health

and welfare effects into perspective.

! Chapters 4 through 7 then assess dosimetry aspects, experimental (controlled human

exposure and laboratory animal) studies, and epidemiologic (field/panel; other observational)

studies. Chapter 8 then provides an integrative synthesis of key findings and conclusions

derived from the preceding chapters with regard to ambient O3 concentrations, human

exposures, dosimetry, and health effects.

! Chapter 9 deals with effects of O3 on vegetation, crops, and natural ecosystems, whereas

Chapter 10 evaluates tropospheric O3 relationships to alterations in surface-level UVB flux

and climate change and Chapter 11 assesses materials damage (these all being key types of

welfare effects of relevance to decisions regarding secondary O3 NAAQS review). 

E.2 ATMOSPHERIC CHEMISTRY AND PHYSICS OF TROPOSPHERIC

OZONE FORMATION

Key findings/conclusions from Chapter 2 regarding the chemistry and physics of surfacelevel O3 formation include the following:
E-4

! Ozone (O3) is a secondary pollutant formed by atmospheric reactions involving two classes

of precursor compounds, volatile organic compounds (VOCs) and nitrogen oxides (NOx). 

Carbon monoxide also contributes to O3 formation. 

! The formation of O3 and associated compounds is a complex, nonlinear function of many

factors, including the intensity and spectral distribution of sunlight; atmospheric mixing and

other atmospheric processes; and the concentrations of precursors in ambient air. 

! The photochemical oxidation of almost all anthropogenic and biogenic VOCs is initiated by

reaction with hydroxyl (OH) radicals. At night, when they are most abundant, NO3 radicals

oxidize alkenes. In coastal and other select environments, Cl and Br radicals can also initiate

the oxidation of VOCs.

! In urban areas, basically all classes of VOCs (alkanes, alkenes, aromatic hydrocarbons,

carbonyl compounds, etc.) and CO are important for ozone formation. Although knowledge

of the oxidative mechanisms of VOCs has improved in recent years, gaps in knowledge

involving key classes, such as aromatic hydrocarbons, still remain. For example, only about

half of the carbon initially present in aromatic hydrocarbons in smog chamber studies form

compounds that have been identified. 

! In addition to gas phase reactions, other reactions also occur on the surfaces of or within

cloud droplets and airborne particles. Most of the well-established multiphase reactions tend

to reduce the rate of O3 formation in polluted environments. Direct reactions of O3 and

atmospheric particles appear to be too slow to reduce O3 formation significantly at typical

ambient PM levels. 

! Oxidants other than O3 are found in the gas phase and in particles. The chemistry occurring

in particle bound-water and, hence, the mechanisms leading to the formation of reactive

oxygen species in particles are largely unknown. 
E-5

! Organic hydroperoxides produced in the oxidation of monoterpenes by O3 could contribute

substantially to secondary organic aerosol formation.

! Our basic understanding of meteorological processes associated with summertime O3

episodes has not changed over recent years. However, the realization is growing that longrange transport processes are important for determining O3 concentrations at the surface. In

addition to synoptic scale flow fields, nocturnal low-level jets can transport pollutants

hundreds of km from their sources in either the upper boundary layer or the lower free

troposphere. Turbulence then brings O3 and other pollutants to the surface.

! Even in the absence of photochemical reactions in the troposphere, some O3 would be found

near the earth’s surface due to its downward transport from the stratosphere. Intrusions of

stratospheric O3 that reach the surface are rare. Much more common are intrusions that

penetrate to the middle and upper troposphere. However, O3 transported to the middle and

upper troposphere can still affect surface concentrations through various mechanisms that

mix air between the planetary boundary layer and the free troposphere above.

! Associations between daily maximum O3 concentration and temperature vary across the

United States and depend on location. In some areas (e.g., Baltimore, MD and surrounding

areas), there is a strong positive association. In other areas (e.g., Phoenix, AZ), there is little

association.

! Chemistry transport models are used to improve understanding of atmospheric chemical and

physical processes, as well as to develop air pollution control strategies. Model evaluation

does not merely involve a straightforward comparison between model predictions and

observed concentration fields of a pollutant of interest (e.g., O3). Such comparisons may not

be meaningful because it is difficult to determine if agreement between measurements and

model predictions truly represents an accurate treatment of physical and chemical processes

in the model or the effects of compensating errors in model routines.
E-6

! The main methods currently used for routine monitoring of ambient ozone are based on

chemiluminescence or UV absorption. Measurements at most ambient monitoring sites are

based on UV absorption. Both of these methods are subject to interference by other

atmospheric components. Studies conducted in Mexico City and in a smog chamber have

found positive interference, but studies conducted in urban plumes did not find evidence for

significant positive interference in the UV absorption technique.

E.3 ENVIRONMENTAL DISPERSAL, AMBIENT CONCENTRATIONS,

AND HUMAN EXPOSURE TO OZONE

Key findings/conclusions derived from Chapter 3 with regard to ambient O3 concentrations

and human exposure are as follows: 

! Ozone is monitored in populated areas in the United States during “ozone seasons,” which

vary in length depending on location. All monitors should be operational from May to

September. However, in many areas, O3 is monitored throughout the year.

! The median of the mean daily maximum 8-h average O3 concentration from May to

September 2000 to 2004 across the U.S. was 0.049 ppm on a countywide average basis. 

Ninety five per cent of countywide mean daily maximum 8-h average O3 concentrations were

less than 0.057 ppm for the same period. Because most monitors are located in the East,

these values should not be taken to represent conditions across the country. 

! The daily maximum 1-h O3 concentrations tend to be much higher in large urban areas or in

areas downwind of large urban areas. For example, daily maximum 1-h O3 concentrations in

Houston, TX approached 0.20 ppm during the same period. 

! Daily maximum 8-h average O3 concentrations are lower than, but are highly correlated

with, 1-h daily maximum O3 concentrations. For example, in the Baltimore, MD area, the

correlation coefficient between the two quantities was 0.98 for data obtained from May to

September 1994 to 2004.
E-7

! Within individual metropolitan statistical areas (MSAs), O3 tends to be well correlated across

monitoring sites. However, there can be substantial spatial variations in concentrations. 

Ozone in city centers tends to be lower than in regions either upwind or downwind of the

center, because of titration by NO emitted by motor vehicles. 

! Ozone concentrations tend to peak in early- to mid-afternoon in areas where there is strong

photochemical production and later in the day in areas where transport is more important in

determining O3 abundance. 

! Summertime maxima in O3 concentrations occur in areas in the United States where there is

substantial photochemical activity involving O3 precursors emitted from human activities. 

Maxima can occur anytime from June through August. 

! Springtime maxima are observed in relatively remote sites in the western United States and

at various other relatively unpolluted sites throughout the Northern Hemisphere. Relatively

high O3 concentrations can also be found during winter in several cities throughout the

southern United States. 

! Long-term trends in O3 concentrations reflect notable decreases over time throughout the

United States, with decreases nationwide of approximately 29% in 2nd highest 1-h O3

concentrations from 1980 to 2003 and of about 21% in 4th highest 8-h O3 concentrations

during the same time period. 

! These trends include dramatic decreases from peak 1-h O3 levels of 0.4 to 0.6 ppm seen in

the Los Angeles area at times in the late 1950's to 1970's to current peak levels of 0.17 ppm

and 0.15 ppm (1-h and 8-h avg, respectively) seen in the Los Angeles basin during

2000-2003. 

! Downward trends in the upper tail of the O3 concentration distribution do not necessarily

reflect trends for O3 values towards the center of the O3 concentration distribution at national

parks. Concentrations toward the center of the distribution have remained more or less
E-8

constant, and O3 values in the lower tail of the distribution show some evidence of slight

increases on a nationwide basis. 

! Policy relevant background (PRB) O3 concentrations are used for assessing risks to human

health associated with O3 produced from anthropogenic sources in the United States, Canada

and Mexico. Because of the nature of the definition of PRB concentrations, they cannot be

derived from observations directly, instead they must be derived from model estimates. 

! Current model estimates indicate that PRB O3 concentrations in the United States surface air

are generally 0.015 ppm to 0.035 ppm. Such concentrations decline from spring to summer

and are generally <0.025 ppm under conditions conducive to high O3 episodes. PRB O3

concentrations may be higher, especially at high altitude sites during the spring, due to

enhanced contributions from (a) pollution sources inside and outside North America and

(b) stratospheric O3 exchange. 

! Only one model (GEOS-Chem) is documented in the literature for calculating PRB O3

concentrations. Estimated PRB O3 values are likely 10 ppbv too high in the Southeast in

summer and are accurate within 5 ppbv in other regions and seasons.

! Sufficient data for other oxidants (e.g., H2O2, PAN) and oxidation products (e.g., HNO3,

H2SO4) in the atmosphere are not available for use in epidemiologic time series studies. 

Limited data for oxidants besides O3 in the gas and particle phases suggest that their

combined concentrations are probably <10 % that of O3. 

! Relationships between O3 and PM2.5 are complex, in part because PM is not a distinct

chemical species, but is a mix of primary and secondary species. For example, PM2.5

concentrations were positively correlated with O3 during summer, but negatively correlated

with O3 during the winter at Ft. Meade, MD. Similar relationships were found for PM10 and

O3 in data collected in a number of urban areas during the 1980s. 
E-9

! Humans are exposed to O3 either outdoors or in various microenvironments. Ozone in

indoor environments results mainly from infiltration from outdoors. Once indoors, O3 is

removed by deposition on and reaction with surfaces and reactions with other pollutants. 

Hence, O3 levels indoors tend to be notably lower than outdoor O3 concentrations measured

at nearby monitoring sites, although the indoor and ambient O3 concentrations tend to vary

together (i.e., the higher the ambient, the higher the indoor O3 levels). 

! Personal exposure to O3 tends to be positively associated with time spent outdoors. 

Although O3 concentrations obtained at stationary monitoring sites may not explain the

variance in individual personal exposures, they appear to serve reasonably well as surrogate

measures for aggregate personal exposures. 

! Atmospheric reactions between O3 and certain other ambient airborne contaminants, e.g.,

terpenes emitted by vegetation or wood products, contribute to generation of ultrafine

particles, with formation of such particles being observed in both urban and rural areas. 

These reactions also occur in indoor environments and involve O3 infiltrating from outdoors

and terpenes emitted by household products (e.g., air fresheners). Gaseous products

resulting from such reactions may also be toxic. 

E.4 OZONE DOSIMETRY AND HEALTH EFFECTS

This section summarizes the main conclusions derived from the integrated synthesis of

information regarding health effects associated with ambient O3 exposures. The conclusions are

based on O3 dosimetry evaluations and human clinical, animal toxicologic, and epidemiologic

studies which have evaluated health effects associated with short-term, repeated, and long-term

exposures to O3 alone or in combination with other ambient pollutants. The controlled human

exposure (or “clinical”) studies provide the clearest and most compelling evidence for human

health effects directly attributable to acute exposures to O3 per se. The evidence from human

and animal toxicologic studies presented in Chapters 4, 5, and 6 are further useful in not only

providing insights into possible mechanisms of action underlying different types of O3-related

health effects but, also, in helping to provide biological plausibility for health effects observed in
E-10

epidemiologic studies assessed in Chapter 7. The studies have also been useful in identifying

susceptible and vulnerable populations that are at potentially greater risk for effects of O3

exposure. Overall, the new findings generally support and build further upon key health-related

conclusions drawn in the previous 1996 AQCD, as summarized below. 

1. Dosimetric Considerations

Chapter 4 discusses dosimetric issues, including factors that are important to consider in

attempting animal-to-human extrapolations of experimentally-induced O3 effects. 

! Dosimetric studies seek to quantify dose and factors affecting the dose of O3 and/or its active

metabolites at specific lung regions, target tissues, or cells. In both humans and animals, the

efficiency of O3 uptake is greater in the nasal passages than the oral pathway. In the lower

respiratory tract, increasing tidal volume increases O3 uptake, whereas increasing flow or

breathing frequency decreases O3 uptake. However, O3-induced rapid shallow breathing

appears to protect the large conducting airways while producing a more even distribution of

injury to the terminal bronchioles.

! In adult human females relative to males, the smaller airways and associated larger surfaceto-volume ratio enhance local O3 uptake and cause somewhat reduced penetration of O3 into

the distal lung. However, it is not clear from these findings if the actual anatomical location

of O3 uptake differs between males and females. 

! Similarly exposed individuals vary in the amount of actual dose received, but O3 uptake is

not predictive of intersubject variability in measures of pulmonary function. 

! The efficiency of O3 uptake is chemical-reaction rate dependent and the reaction products

(hydrogen peroxide, aldehydes, and hydroxyhydroperoxides) created by ozonolysis of lipids

in epithelial lining fluid (ELF) and cell membranes appear to mediate O3 toxicity. 

! Ozone uptake in humans is increased by exposure to NO2 and SO2 and decreased during

the O3 exposure. This suggests that an inflammatory response during exposure to NO2 and
E-11

SO2 may elicit increased production of O3-reactive substrates in the epithelial lining fluid and

that these substrates are depleted by O3 exposure but not by NO2 and SO2 exposures. 

! Prior modeling studies have suggested the proximal alveolar and centriacinar regions as

principal target sites of acute O3-induced cell injury. New experimental work in rats

suggests that the conducting airways are also a primary site of injury. 

! In most clinical studies, humans are exposed to O3 during exercise. Under these conditions,

the switch from nasal to oral breathing, coupled with increases in respiratory flow (as occurs

during exercise), causes a shift in the O3 dose distribution, allowing O3 to penetrate deeper

into the lung and thereby increasing the potential for bronchiolar and alveolar damage.

! Comparisons of acute exposures in rats and humans suggest that, though both species have

similar qualitative responses to O3 exposure, there are interspecies mechanistic disparities

that necessitate careful comparisons of dose-response relationships. Currently available data

suggest that lowest observable effect levels in resting rats are approximately 4- to 5-fold

higher than for exercising humans for some toxicological endpoints, e.g., increases in

bronchoalveolar lavage (BAL) protein or neutrophil (PMN) levels (indicators of O3-induced

lung inflammation responses). 

! Thus, a number of variables seem to affect O3 uptake, notably including route of breathing,

breathing pattern, gender, copollutants, and certain pre-exposure conditions. These

differences are important in order to interrelate experimentally-demonstrated

pathophysiological effects and epidemiologically-observed associations between ambient O3

concentrations and health risks among human population groups.

2. Health Effects of Short-term Exposures to Ozone

The 1996 O3 AQCD assessed a substantial body of evidence from toxicologic, human

clinical, and epidemiologic studies. That AQCD concluded that short-term ambient O3 exposure

resulted in various respiratory health effects, including lung function decrements and increased

respiratory symptoms in both healthy and asthmatic individuals exposed during moderate to
E-12

heavy exercise to O3 concentrations ranging down to the lowest levels (0.12 ppm for 1 h;

0.08 ppm for 6.6 to 8 h) tested in the available controlled human exposure studies. Such

experimentally demonstrated effects were consistent with and lent plausibility to epidemiologic

observations highlighted in the 1996 AQCD of increases in daily hospital admissions and ED

visits for respiratory causes. Epidemiologic evidence also provided suggestive evidence for an

association between short-term O3 exposure and mortality. However, there was essentially no

evidence available in the 1996 O3 AQCD regarding potential cardiovascular effects of short-term

O3 exposure. The newly-available evidence assessed in this revised O3 AQCD notably enhances

our understanding of short-term O3 exposure effects, as summarized below, first in relation to

respiratory morbidity endpoints and then cardiovascular effects and, lastly, mortality. 

A. Respiratory Morbidity

Lung Function: 

! Controlled exposure studies clearly demonstrate acute reversible decrements in lung function

in healthy adults exposed to $0.08 ppm O3 when minute ventilation and/or duration of

exposure are increased sufficiently. On average, spirometric responses to O3 exposure

appear to decline with increasing age starting at approximately 18 to 20 years of age. 

! There is considerable variability in responses between similarly exposed individuals, such

that some may experience distinctly larger effects even when small group mean responses

are observed. For example, healthy adults exposed to 0.08 ppm O3 for 6.6 h with moderate

exercise exhibited a group mean O3-induced decrement in FEV1 of about 6%, but a

decrement of >10% was seen in 23% of these individuals. Also, exposure to 0.06 ppm O3

caused >10% lung function decline in a small percentage (7%) of the subjects. 

! Summer camp field studies conducted in southern Ontario, Canada, in the northeastern U.S.,

and in southern California have also reported lung function responses in pre-adolescent

children associated with ambient O3 levels.

! Repeated acute (1- to 6-h) O3 exposures at 0.12 to 0.45 ppm over several days in controlled

exposure studies typically find that FEV1 response to O3 is enhanced on the second of several
E-13

days of exposure, but spirometric responses become attenuated on subsequent days with

these repeated exposures. However, this tolerance is lost after about a week without

exposure. 

! Animal toxicologic studies also provide extensive evidence that acute O3 exposures alter

breathing patterns so as to cause rapid shallow breathing (i.e., increased frequency and

decreased tidal volume), an effect which attenuates after several days of exposure. 

! Results from controlled human exposure studies and animal toxicologic studies provide clear

evidence of causality for the associations observed between acute (#24 h) O3 exposure and

relatively small, but statistically significant declines in lung function observed in numerous

recent epidemiologic studies. Declines in lung function are particularly noted in children,

asthmatics, and adults who work or exercise outdoors.

Respiratory Symptoms: 

! Young healthy adult subjects exposed in clinical studies to O3 concentrations $0.08 ppm for

6 to 8 h during moderate exercise exhibit symptoms of cough and pain on deep inspiration. 

An increase in the incidence of cough has been found in clinical studies as low as 0.12 ppm

in healthy adults during 1 to 3 h with very heavy exercise and other respiratory symptoms,

such as pain on deep inspiration and shortness of breath, have been observed at 0.16 ppm to

0.18 ppm with heavy and very heavy exercise. These O3-induced respiratory symptoms

gradually decrease in adults with increasing age. With repeated O3 exposures over several

days, respiratory symptoms become attenuated, but this tolerance is lost after about a week

without exposure. 

! The epidemiologic evidence shows significant associations between acute exposure to

ambient O3 and increases in a wide variety of respiratory symptoms (e.g., cough, wheeze,

production of phlegm, and shortness of breath) in asthmatic children. Epidemiologic studies

also indicate that acute O3 exposure is likely associated with increased asthma medication

use in asthmatic children. 
E-14

! On the other hand, an effect of acute O3 exposure on respiratory symptoms in healthy

children is not as clearly indicated by epidemiology studies, consistent with diminished

symptom responses seen in healthy children in human clinical studies.

Airway Inflammation: 

! Inflammatory responses have been observed subsequent to 6.6 h O3 exposures to the lowest

tested level of 0.08 ppm in healthy human adults. Some studies suggest that inflammatory

responses may be detected in some individuals following O3 exposures even in the absence

of O3-induced pulmonary function decrements in those subjects. 

! Repeated O3 exposures over several days leads to an attenuation of most inflammatory

markers. However, none of the several markers of lung injury and permeability evaluated

show attenuation, indicating continued lung tissue damage during repeated exposure. 

! Animal toxicologic studies provide extensive evidence that acute (1 to 3 h) O3 exposures as

low as 0.1 to 0.5 ppm can cause (1) lung inflammatory responses (typified by increased

reactive oxygen species, inflammatory cytokines, influx of PMNs, and activation of alveolar

macrophages); (2) damage to epithelial airway tissues, (3) increases in permeability of both

lung endothelium and epithelium, and (4) increases in susceptibility to infectious diseases

due to modulation of lung host defenses. 

! Consistent with these experimental findings, there is also limited epidemiologic evidence

showing an association between acute ambient O3 exposure and airway inflammation in

children acutely exposed to ambient O3 concentrations (1-h max O3 of approximately

0.1 ppm). 

! The extensive human clinical and animal toxicological evidence, together with the limited

available epidemiologic evidence, is clearly indicative of a causal role for O3 in

inflammatory responses in the airways.
E-15

Airway Responsiveness: 

! Controlled human exposure studies have found that acute O3 exposure causes an increase in

nonspecific airway responsiveness, as indicated by reductions in concentrations of

methacholine or histamine required to produce a given decrease in FEV1 or increase in SRaw.

! Acute (2- or 3-h) O3 exposure at 0.25 or 0.4 ppm of allergic asthmatic subjects, who

characteristically already have somewhat increased airway responsiveness at baseline, was

found to cause further increases in airway responsiveness in response to allergen challenges. 

Also, repeated daily exposure to 0.125 ppm O3 for 4 days exacerbated lung function

decrements in response to bronchial allergen challenges among persons with preexisting

allergic airway disease, with or without asthma. 

! Ozone-induced exacerbation of airway responsiveness persists longer and attenuates more

slowly than O3-induced pulmonary function decrements and respiratory symptom responses. 

Heightened airway responsiveness (reactivity) has also been observed in several laboratory

animal species with acute exposures (1 to 3 h) to 0.5 to 1.0 ppm O3. Ozone increases airway

hyperreactivity to bronchoconstrictive agents (e.g., ovalbumin), and there is a temporal

relationship between inflammatory cell influx and O3-induced increases in airway reactivity. 

Several studies of sensitized laboratory animals showing O3-induced increases in airway

hyperreactivity are consistent with O3 exacerbation of airway hyperresponsiveness reported in

atopic humans with asthma. 

! Airway responsiveness has not been widely examined in epidemiologic studies. However, the

evidence from human clinical and animal toxicological studies clearly indicate that acute

exposure to O3 can induce airway hyperreactivity, thus likely placing atopic asthmatics at

greater risk for more prolonged bouts of breathing difficulties due to airway constriction in

response to various airborne allergens or other triggering stimuli. 

Respiratory Hospital Admissions and Emergency Department Visits: 

! Aggregate population time-series studies observed that ambient O3 concentrations are

positively and robustly associated with respiratory-related hospitalizations and asthma ED
E-16

visits during the warm season. These observations are strongly supported by the human

clinical, animal toxicologic, and epidemiologic evidence for lung function decrements,

increased respiratory symptoms, airway inflammation, and airway hyperreactivity. 

! Taken together, the overall evidence supports a causal relationship between acute ambient O3

exposures and increased respiratory morbidity outcomes resulting in increased ED visits and

hospitalizations during the warm season.

B. Cardiovascular Morbidity

At the time of the 1996 O3 AQCD, the possibility of O3-induced cardiovascular effects was

a largely unrecognized issue. Newly-available evidence has emerged since then which provides

considerable plausibility for how O3 exposure could exert cardiovascular impacts. 

! Direct O3 effects such as O3-induced release from lung epithelial cells of platelet activating

factor (PAF) that may contribute to blood clot formation that would increase the risk of

serious cardiovascular outcomes (e..g, heart attack, stroke, mortality). Also, interactions of

O3 with surfactant components in epithelial lining fluid of the lung results in production of

oxysterols and reactive oxygen species that may exhibit PAF-like activity contributing to

clotting and/or exert cytotoxic effects on lung and heart cells. 

! Indirect effects of O3 may involve O3-induced secretions of vasoconstrictive substances

and/or effects on neuronal reflexes that may result in increased arterial blood pressure and/or

altered electrophysiologic control of heart rate or rhythm. Some animal toxicological studies

have shown O3-induced decreases in heart rate, mean arterial pressure, and core temperature.

! Some field/panel studies that examined associations between O3 and various cardiac

physiologic endpoints have yielded limited epidemiologic evidence suggestive of a potential

association between acute O3 exposure and altered HRV, ventricular arrhythmias, and

incidence of MI. 

! Highly suggestive evidence for O3-induced cardiovascular effects is provided by a few

population studies of cardiovascular hospital admissions which reported positive O3
E-17

associations during the warm season between ambient O3 concentrations and cardiovascular

hospitalizations. Only one controlled human exposure study that evaluated effects of O3

exposure on cardiovascular health outcomes found no significant O3-induced differences in

ECG, heart rate, or blood pressure in healthy or hypertensive subjects, but did observe an

overall increase in myocardial work and impairment in pulmonary gas exchange. 

! Overall, this generally limited body of evidence is highly suggestive that O3 directly and/or

indirectly contributes to cardiovascular-related morbidity, but much remains to be done to

more fully substantiate links between ambient O3 exposure and adverse cardiovascular

outcomes. 

C. Mortality

Numerous recent epidemiologic studies conducted in the United States and abroad have

investigated the association between acute exposure to O3 and mortality. Results from several

large U.S. multicity studies as well as several single-city studies indicate a positive association

between increases in ambient O3 levels and excess risk of all-cause (nonaccidental) daily

mortality. 

! Consistent with observed O3-related increases in respiratory- and cardiovascular-related

morbidity, several newer multicity studies, single-city studies, and several meta-analyses of

these studies have provided relatively strong epidemiologic evidence for associations

between short-term O3 exposure and all-cause mortality, even after adjustment for the

influence of season and PM. 

! Determining cause-specific mortality is more difficult due to reduced statistical power by

which to examine cause-specific associations and the lack of clarifying information on

contributing causes of death. That is, attribution to one or the other of the more specific

cardiopulmonary causes may underplay contributions of chronic cardiovascular disease to

“respiratory” deaths (e.g., a heart attack victim succumbing to acute pneumonia) or vice

versa. 
E-18

! Consistently positive associations have been reported for O3-related cardiovascular mortality

across approximately 30 studies, with two well-conducted multicity studies in the United

States and Europe yielding small, but statistically significant positive associations. 

! Both animal and human studies provide evidence suggestive of plausible pathways by which

risk of respiratory or cardiovascular morbidity and mortality could be increased by ambient

O3 either acting alone or in combination with copollutants in ambient air mixes. 

! This overall body of evidence is highly suggestive that O3 directly or indirectly contributes to

non-accidental and cardiopulmonary-related mortality, but additional research is needed to

more fully establish underlying mechanisms by which such effects occur.

3. Health Effects of Long-term Exposures to Ozone

In the 1996 O3 AQCD, the available epidemiologic data provided only suggestive evidence

that respiratory health effects were associated with chronic O3 exposure. Animal toxicologic

studies indicated that chronic O3 exposure caused structural changes in the respiratory tract, and

simulated seasonal exposure studies in animals suggested that such exposures might have 

cumulative impacts. As summarized below, recent studies are generally consistent with the

conclusions drawn in the previous 1996 AQCD.

A. Respiratory Morbidity

Lung Function: 

! Recent epidemiologic studies observed that reduced lung function growth in children was

associated with seasonal exposure to O3; however, cohort studies investigating the effect of

annual or multiyear O3 exposure observed little clear evidence for impacts of longer-term,

relatively low-level O3 exposure on lung function development in children. 

! The epidemiologic data, collectively, indicate that the current evidence is suggestive but

inconclusive for respiratory health effects from long-term O3 exposure.
E-19

Morphological Changes: 

! Animal toxicologic studies continue to show chronic O3-induced structural alterations in

several regions of the respiratory tract including the centracinar region. Morphologic

evidence from some recent studies using exposure regimens that mimic seasonal exposure

patterns report increased lung injury compared to conventional chronic stable exposures. 

! Infant rhesus monkeys repeatedly exposed to 0.5 ppm 8h/day O3 for 11 episodes exhibited:

(1) remodeling of the distal airways; (2) abnormalities in tracheal basement membrane; (3)

eosinophil accumulation in conducting airways; and (4) decrements in airway innervation. 

Long-term O3 exposure of rats to 0.5 or 1.0 ppm for 20 months resulted in upper respiratory

tract mucus metaplasia and hyperplasia in the nasal epithelium (0.25 or 0.5 ppm, 8h/day,

7days/wk for 13 weeks). 

! The persistent nature of these cytological changes raise the possibility of long-lasting

alterations in human airways in response to chronic O3 exposure, but it is highly uncertain as

to what long-term patterns of exposure or O3 concentrations in humans may be requisite to

produce analogous morphological changes. Nor is it now possible to characterize the

possible magnitude or severity of any such effects occurring in humans in response to

ambient O3 exposures at levels observed in the United States. 

Incidence of Lung Cancer: 

! The weight of evidence from recent animal toxicological studies and a very limited number

of epidemiologic studies do not support ambient O3 as a pulmonary carcinogen.

B. Mortality

! Results from the few available epidemiologic studies are inconsistent regarding the

association between long-term exposure to O3 and mortality. There is little evidence to

suggest a causal relationship between chronic O3 exposure and increased risk for mortality in

humans.
E-20

4. Health Effects of Ozone-Containing Pollutant Mixtures

The potential interaction of pollutant mixtures with O3 is poorly understood and the animal

studies reviewed in the 1996 O3 AQCD reported additive, synergistic or antagonistic effects

depending on the exposure regimen and the endpoint studied. A few new controlled human

exposure and animal toxicology studies reviewed in Chapters 4, 5, and 6 investigated health

effects associated with O3-containing pollutant mixtures of near ambient levels. As noted below,

recent studies, although generally consistent with conclusions drawn in the 1996 O3 AQCD, have

added some new information, particularly with regard to interactions between O3 and PM.

! Controlled human exposure studies indicate that continuous exposure of healthy human

adults to SO2 or NO2 increases bolus dose O3 absorption, suggesting that co-exposure to

other gaseous pollutants in the ambient air may enhance O3 absorption. 

! Other controlled human exposure studies that evaluated response to allergens in asthmatics

(allergic and dust-mite sensitive) suggest that O3 enhances response to allergen challenge. 

Consistent with these findings, animal toxicology studies also reported enhanced response to

allergen on exposure to O3.

! A few other animal toxicology studies that exclusively investigated the co-exposure of PM

and O3 reported increased response (lung tissue injury, inflammatory and phagocytosis) to

the mixture of PM + O3 compared to either PM or O3 alone. 

! Recent investigations on the copollutant interactions using simulated urban photochemical

oxidant mixes suggest the need for similar studies in understanding the biological basis for

air pollutant mixture effects observed in epidemiologic studies.

5. Susceptibility or Vulnerability to Effects Associated with Exposure to Ozone

Various factors have been shown to influence individuals’ responses to environmental air

pollutants. Factors that increase susceptibility to O3-related effects include innate factors, such

as genetic predisposition or developmental effects, or disease status. Other factors can lead to

enhanced vulnerability to O3-related effects, such as heightened exposures or activity patterns. 

In the 1996 O3 AQCD, available evidence suggested that children, asthmatics, and outdoor
E-21

workers were populations that may be more susceptible or vulnerable to effects of O3 exposure. 

In addition, controlled human exposure studies also demonstrated a large variation in sensitivity

and responsiveness to O3 in studies of healthy subjects, but the specific factors that contributed

to this intersubject variability were yet to be identified. Recent studies have built upon the

evidence available in the previous review. Factors related to susceptibility or vulnerability to O3

exposure-related effects are briefly summarized below:

People with Preexisting Pulmonary Diseases: 

! Ozone-induced differential responses in lung function and AHR in people with allergic

rhinitis suggest that asthmatics have potentially greater responses than healthy people with

exposure to O3. There is a tendency for slightly increased spirometric responses in mild

asthmatics and allergic rhinitics relative to healthy young adults. Spirometric responses in

asthmatics appear to be affected by baseline lung function, i.e., responses increase with

disease severity. 

! Repeated O3 exposure over several days has been shown to increase responsiveness to

bronchial allergen challenge in subjects with preexisting allergic airway disease, with or

without asthma. Asthmatics also show a significantly greater neutrophil response (18 h

postexposure) than similarly-exposed healthy individuals. 

! Epidemiologic studies have reported associations with a range of respiratory health outcomes

in asthmatics, from decreases in lung function to hospitalization or ED visits for asthma, thus

supporting this population group as being likely to experience increased risk for O3-induced

health effects. 

! Controlled human exposure studies have not found evidence of larger spirometric changes in

people with COPD relative to healthy subjects, this may be due to the fact that most people

with COPD are older adults who would not be expected to have such changes based on their

age. However, new epidemiologic evidence indicates that people with COPD may be more

likely to experience other effects, including emergency room visits, hospital admissions, or

premature mortality.
E-22

Age-related: 

! Controlled human exposure studies have shown that lung function responses to O3 varies

with age, with responsiveness generally diminishing after about 18 to 20 years of age. 

Children and older adults thus have lesser respiratory symptoms with O3 exposure than

young healthy adults. Potentially increased O3 doses can be received by individuals

experiencing less severe respiratory symptoms. 

! Evidence from newer epidemiologic studies supports the 1996 O3 AQCD conclusions that

children are more likely at increased risk for O3-induced health effects. Notably,

epidemiologic studies have indicated adverse respiratory health outcomes associated with O3

exposure in children. In addition, recently published epidemiologic studies also suggest that

older adults (aged $65 years) appear to be at excess risk of O3-related mortality or

hospitalization. 

Heightened vulnerability due to greater exposures: 

! Epidemiologic studies have provided some evidence to indicate that outdoor workers are

more vulnerable to O3-related effects, which is likely related to their increased exposure to

ambient air pollution. 

! Controlled human exposure studies clearly established differential biological response to O3

based on physical activity (exertion). Epidemiologic studies also suggest that exercising

(moderate to high physical exertion) children and adolescents appear to demonstrate

increased responsiveness to ambient concentrations of O3 and may be more likely to

experience O3-induced health effects. Animal studies show a similar impact of exercise on

responsiveness to O3.

Genetic susceptibility: 

! Animal toxicologic studies provide supportive evidence to the observations of innate

susceptibility. Various strains of mice and rats have demonstrated the importance, in

general, of genetic background in O3 susceptibility. Moreover, genetic and molecular
E-23

characterization studies in laboratory animals identified genetic loci responsible for both

sensitivity and resistance. 

! New human clinical and epidemiologic studies also have shown that genetic polymorphisms

for antioxidant enzymes and inflammatory genes (GSTM1, NQO1, and Tnf-") may modulate

the effect of O3 exposure on pulmonary function and airway inflammation. 

E.5 VEGETATION AND ECOLOGICAL EFFECTS

Data published since 1996, as assessed in Chapter 9 and associated annex materials,

continues to support and strengthen the conclusions of previous O3 AQCDs. The main

findings/conclusions derived from the current Chapter 9 assessment of O3 ecological effects are

as follows.

General

! The ecological effects of O3 appear to be widespread across the United States based on recent

biomonitoring studies using clover and other species grown in plots across the United States,

as well as regional forest health visible injury surveys.

! Some plant community compositions may be shifting based on recent studies of competition

among plants in managed pasture lands, as well in natural unmanaged lands where increased

O3 effects on sensitive species in the community can reduce their presence in the community. 

! Research to date has focused at the species level, with very few studies at the ecosystem

level. The lack of data at this organizational level hampers the assessment of O3 risk to

ecosystem services, such as water quality and quantity, that contribute to human well-being.

Methodologies

! New methodologies coming into use since 1996 have not fundamentally altered our

understanding of O3 effects on plants or the conclusions of the 1996 O3 AQCD. Since 1996,

there has been a shift from chamber-based studies to the field-based approach, including plot
E-24

and regional visible injury surveys and the use of the non-chambered free air CO2 exposure

(FACE) systems. The FACE system results support earlier observations of foliar injury and

reduced volume growth in aspen and indicate reduced yield in soybean cultivars similar to

earlier studies in open-top chamber systems (OTC). 

! The use of biomonitoring since 1996 has advanced identification and symptom verification

of sensitive species and has been a useful tool for indicating the extent of O3 effects across

most of the eastern and southeastern United States and many parts of the West.

! The development and improvement of stomatal models for predicting O3 uptake in Europe

have fostered more universal measures of exposure response. These simulation tools may

provide a better means to relate ambient exposure to plant response in the future but

currently are insufficient for use across broad geographical areas of the United States. 

! Since 1996, the use of passive samplers for monitoring O3 in rural and remote areas has

expanded, offering a potential for improved exposure data in areas not actively monitored. 

The testing and development of these samplers will ultimately provide a strategy to expand

air quality monitoring into areas for which exposure characterization is currently done by

geospatial extrapolation techniques such as Kriging.

Mode of Action

! There are several steps in the process of O3 uptake and toxicity that are better understood

now than in 1996, based on new information gained in part by use of improved molecular

tools for following rapid changes that occur within the leaf. These advancements are

important for refining hypotheses on O3 uptake and improving understanding of exposureresponse relationships.

! Ozone entrance into the leaf through the stomata remains the critical step in O3 sensitivity. 

Although the initial reactions within the leaf are still unclear, the involvement of H2O2 is

clearly indicated. The initial sites of membrane reactions seem to involve transport
E-25

properties. The primary set of metabolic reactions that O3 triggers currently includes those

typical of “wounding” responses generated by leaf cutting or by insect attack. 

! The alteration of normal metabolism due to wounding spreads outside the cytoplasm. One of

the secondary reactions is linked to a senescence response. The loss of photosynthetic

capacity is linked to lower productivity (although not fully elucidated) and to problems with

efficient translocation of carbon. 

! Chronic O3 effects are linked to the senescence process or some physiological process

closely linked to senescence, e.g., translocation, re-absorption, allocation of nutrients and

carbon.

Modification of Growth Response

! Many biotic and abiotic factors, including insects, pathogens, root microbes and fungi,

temperature, water and nutrient availability, and other air pollutants, as well as elevated CO2,

influence or alter the plant’s response to O3. A few studies published since 1996 have

improved our understanding of the role of these interactions in modifying O3-induced plant

responses.

! Biotic Interactions: Recent studies have supported earlier conclusions that O3 often increases

the likelihood and success of insect attacks, at least by chewing insects. Less is known

regarding sucking insects (e.g., aphids). It seems that some insect problems could be

exacerbated by increased O3 exposure, but predicting any particular O3-insect interaction is

not possible at this time.

! Biotic Interactions: More information is available regarding disease interactions. 

Ozone exposure generally increases plant diseases associated with facultative necrotrophic

plant pathogens. Pathogens that benefit from damage to cells are enhanced by O3 stress to

their hosts, whereas pathogens that require healthy hosts are depressed by O3 stress.
E-26

! Biotic Interactions: A few new studies have demonstrated O3 impacts on intraspecific plant

competition. In grass-legume pastures, the legume component is more O3-sensitive and is

reduced over time. Similarly, grass competition on pine seedlings can enhance O3 effects on

the seedlings, possibly through the grass’s ability to outcompete the seedlings for water.

! Abiotic Interactions: New information on the role of abiotic or physical factors interacting

with O3 stress support 1996 O3 AQCD conclusions. Some studies have shown an increasing

effect of O3 with increasing temperature, but others have shown little effect of temperature. 

Temperature is an important variable affecting plant O3 response in the presence of elevated

CO2 levels associated with climate change. It also appears that low temperatures are

important, in that O3 exposure sensitizes plants to low temperature stress.

! Abiotic Interactions: New information on the role of drought and water availability

published since 1996 confirms earlier conclusions regarding the increased effects of O3

where readily available soil moisture results in increased needle/leaf conductivity and thus

increased O3 uptake. Additional studies demonstrated again the partial “protection” against

adverse effects of O3 by drought. There was also evidence that O3 predisposes plants to

drought stress. The net results of these interactions are negative, at least in the short term,

although longer lived species like trees could benefit from increased water use efficiency.

Effects-Based Exposure Indices

! Exposure indices are metrics relating plant response (i.e., growth or yield) to monitored

ambient O3 concentrations over time to provide a consistent metric for reviewing and

comparing exposure-response effects obtained from various studies. Such metrics may also

provide a basis for developing air quality standards that are protective of ecological

resources. The 1996 O3 AQCD focused on research where a large number of indices were

developed that included various functional and statistical summaries of ambient hourly

concentrations over designated time periods. The development of those indices focused on

considering and including some, but not all, the factors that affect O3 uptake and expression

of effects.
E-27

! Conclusions from the 1996 O3 AQCD regarding an ambient-exposure based index are still

valid. No information since 1996 significantly alters the basic conclusions, and most studies

in this interim have further supported them. The key conclusions are as follows:

! Ozone effects in plants are cumulative;

! Higher O3 concentrations appear to be more important than low concentrations in

eliciting a response;

! Plant sensitivity to O3 varies with time of day and plant development stage; and

! Exposure indices that accumulate the O3 hourly concentrations and preferentially weight

the higher concentrations have a better statistical fit to growth/yield response than do

mean or peak indices.

! Based on the current state of knowledge, exposure indices that differentially weight the

higher hourly average O3 concentrations but include the mid-level values represent the best

approach for relating vegetation effects to O3 exposure in the United States. A large database

for crops and tree seedlings exists and has been used for establishing exposure-response

relationships and predicting effects for a range of exposure concentrations. In 1996, EPA

considered three specific concentration-weighted indices for use as air quality indicators: the

cutoff concentration-weighted SUM06, the AOT60, and the sigmoid-weighted W126. All

three performed equally well based on goodness-of-fit tests. Since 1996, there have been no

published experimental studies that would alter the consideration of these concentrationweighted cumulative indices.

! Studies available since 1996 strengthen earlier conclusions on the role of exposure

components such as duration, concentration, and temporal patterns in determining plant

growth response to O3 exposure. New studies since 1996 have shown experimentally the

disconnection of peak events and maximal stomatal conductance at a variety of sites. The

identification of sensitivity linked to time of day (i.e., period of maximum conductance) was

reported in the 1996 O3 AQCD. The new studies will offer future avenues for building this

temporal component into an exposure index. Similarly, recent reviews of the plant literature

have reported a large number of species with nighttime conductance capable of O3 uptake. 

The designated time interval for cumulating exposure (i.e., 12 to 24 h) needs reconsideration.
E-28

! Recent research in Europe has focused on a flux-based approach to improve upon the

ambient air concentration-based (i.e., exposure indices) approach to assess risk from O3

across different climate regions. However, such approaches need further development to

incorporate the necessary complexity across space and time to be non-site and non-species

specific. Also, at this time, the database is inadequate for linking O3 flux to growth

responses. 

Ozone Exposure-Plant Response Relationships

! Data published since 1996 continues to support and strengthen the conclusions of previous

O3 AQCDs that there is strong evidence that current ambient O3 concentrations cause

(1) decreased growth and biomass accumulation in annual, perennial and woody plants,

including agronomic crops, annuals, shrubs, grasses, and trees; (2) decreased yield and/or

nutritive quality in a large number of agronomic and forage crops; and (3) impaired aesthetic

quality of many native plants and trees by increased foliar injury.

! Since 1996, the published studies have supported earlier conclusions on reduced growth and

yield in a number of crops and trees, and have used multiple approaches, including regional

visible injury surveys, measured growth responses across ambient exposure gradients,

empirical exposure studies in chambered and non-chambered systems, as well as process

model simulations.

! Studies of growth response using open-top chambers have provided useful data for assessing

O3 impact on common and economically valuable species, and developing functional

growth-response models that enable the prediction of O3 impact over a wide range of

ambient air exposures. The studies were designed to maximize statistical robustness by

replicating a number of treatments and, at the same time, considering issues of extrapolation

by conducting the studies across a wide range of crop-growing regions and forested sites in

the United States. Such designs allowed the studies to account for climate and growing

conditions, as well as regional crop growing practices.
E-29

! Recent exposure studies in Illinois using the non-chambered FACE system with soybean

cultivars reported reductions in yield of soybeans in two year-long studies similar to the

reductions found in multiple soybean studies conducted in the 1980’s using open-top

chamber systems. Multiple-year exposure studies using the FACE system (a) found foliar

injury and reduced volume growth in aspen and maple similar to results reported from earlier

open-top chamber studies and (b) highlight the importance of multiyear studies with

longer-growing species.

! Since 1996, additional studies have supported 1996 O3 AQCD conclusions that deciduous

trees are generally less O3 sensitive than most annual species or crop plants, with the

exception of a few very sensitive genera (e.g., sensitive clones or genotypes of Populus) and

sensitive species (e.g., black cherry). Coniferous species have a wide range of O3

sensitivities but, in general, are less sensitive than deciduous species. Among conifers, the

slower-growing species are less sensitive than faster-growing species. Data from a few

European studies support these conclusions. 

! For all types of perennial vegetation, cumulative effects over more than one growing season

may be important; studies of one or a few seasons may under- or overestimate O3 impacts on

these species. Results from multiyear studies sometimes found a pattern of increased effects

in subsequent years, whereas other studies reported growth decreases due to O3 that become

less significant or disappear over time. It is difficult to conduct empirical experiments with

long-lived trees, because even multiyear exposures only account for a small fraction of the

tree’s lifetime. Model simulations of growth have been a tractable approach to account for

time and changing climate in assessing the impact on long-lived trees. 

Ecosystem Effects

! There is strong evidence that O3, in locations where ambient levels are relatively high, is an

important stressor of ecosystems, with documented impacts on the biotic condition,

ecological processes, and chemical/physical nature of natural ecosystems. Experimentally

documented effects on individual keystone species and their associated microflora and fauna

may cascade through the ecosystem to the landscape level, but this has not been quantified.
E-30

! Systematic injury surveys (e.g., USDA Forest Health Network; Europe’s ICP Forests) show

that foliar injury and crown/canopy deformations occur in O3-sensitive species in many

regions of the United States and Europe. However, the lack of general correspondence

between foliar symptoms and growth effects means that other methods must be used to

estimate regional effects of O3 on tree growth. Regional studies of radial growth in mature

trees, combined with data from many controlled studies with seedlings and a few studies

with mature trees, suggest that ambient O3 may be reducing the growth of mature trees in

some U.S. locations.

! The use of physiological-based process models to simulate tree growth, combined with

stand-level models predicting forest composition and productivity, is an approach being used

recently in assessing O3 impacts on forests. These tools suggest that modest O3 effects on

growth may accumulate over time and interact with effects of other naturally occurring

stresses (e.g., drought, nutrient availability). For mixed-species stands, the models predict

that overall stand growth may not be affected, but competitive interactions among species

may change the composition due to growth reductions in sensitive species.

! The knowledge base for examining the range of ecological effects of O3 on natural

ecosystems is growing, but significant uncertainties remain regarding O3 effects at the

ecosystem level. A number of significant areas for investigation that would improve our

ability to assess O3 effects on ecosystems and the services they provide for human well-being

have been outlined and discussed (see Annex AX-9).

E.6 TROPOSPHERIC OZONE EFFECTS ON UV-B FLUX AND ITS

ROLE IN CLIMATE CHANGE

Molecular properties specific to O3 include a capacity for absorbing incoming ultraviolet

(UV) and infrared (IR) radiation, and both incoming solar and outgoing terrestrial IR radiation. 

Consequently, O3 plays an essential role in shielding the earth’s surface from harmful levels of

UV-B radiation, by way of the stratospheric O3 layer. Its effectiveness as a screen for the

residual UV-B flux that penetrates the stratosphere and passes into the troposphere and its role in
E-31

reducing UV-induced human health effects are addressed in Chapter 10. The radiationabsorbing properties of O3 also make it a greenhouse gas (GHG) having global and regional

consequences for climate, as also addressed in Chapter 10. Important conclusions from

Chapter 10 are summarized below. 

! The distribution of O3 within the atmosphere. Ozone is distributed very unevenly within the

atmosphere, with ~90% of the total atmospheric burden present in the stratosphere. The

remaining ~10% is distributed within the troposphere, with higher relative concentrations

near the source of its precursors at the surface. Concentrations of O3 at the mid- and uppertroposphere vary, depending upon meteorological conditions.

! Multiple factors govern the flux of UV-B radiation at the Earth’s surface. Latitude and

altitude are the two most important factors that define the residual UV-B flux at the surface. 

Natural variation in the total column density of stratospheric O3 is also an important factor. 

All of these factors are followed in importance by tropospheric clouds, particulate matter

(PM) and O3. The effect of natural stratospheric variation, clouds, PM and tropospheric O3

on UV fluxes within the troposphere and at the surface are each very difficult to predict. 

! A UV-B “climatology” is needed to predict human exposure levels. A UV-B climatology,

representing patterns and trends in UV-B flux at the Earth’s surface, must be based on

extended in situ observations in order to adequately capture natural variability and the effects

of human activities on atmospheric UV-B absorbers. At present, the body of UV-B

measurements cannot support the development of a climatology.

! Human exposure to UV-B radiation. Quantitative evaluation of human exposure to UV-B

radiation is necessary to perform health risk assessment for UV-B-related health effects. 

Individuals who participate in outdoor sports and activities, work outdoors, live in

geographic areas with higher solar flux, and/or engage in high-risk behavior (e.g., extended

sun bathing) can reasonably be projected to be at increased risk for higher UV radiation

exposures. However, little is known about the impact of variability in these factors on

individual exposure to UV radiation.
E-32

! Human health effects of UV-B radiation. Exposure to UV-B radiation is associated with

increased risk of erythema, nonmelanoma and melanoma skin cancers, ocular damage, and

immune system suppression. Some studies have attempted to estimate the potential effects

of changes in surface-level UV flux resulting from stratospheric O3 depletion on these health

outcomes; however, the numerous simplifying assumptions made in the assessments limit the

usefulness of the risk estimates. The effect of changes in surface-level O3 concentrations on

UV-induced health outcomes cannot yet be critically assessed within reasonable uncertainty.

! Vitamin D-related health benefits of UV-B radiation. A potential health benefit of increased

UV-B exposure relates to the production of vitamin D in humans. Several studies have

found that UV-B radiation, by increasing vitamin D production, is associated with reduced

risks of various cancers. However, as with other impacts of UV-B on human health, this

beneficial effect of UV-B has not been studied in sufficient detail to allow for a credible

health benefits assessment. No study has been done of the decreased risk of cancer resulting

from increased UV radiation attributable to decreased tropospheric O3 levels, but the change

in risk is expected to be unappreciable.

! Ozone is a potent GHG. Ozone traps incoming solar radiation at both ends of the spectrum,

as well as shortwave radiation that is scattered from high-albedo portions of the Earth’s

surface. Outgoing terrestrial IR is absorbed by O3 within the range where water vapor does

not absorb, so that natural variability in humidity does not alter its radiative impact. These

effects directly force climate. By participating in the oxidative chemistry of the atmosphere,

O3 can indirectly and negatively force climate by the removal of other greenhouse gases.

! Multiple factors influence the forcing effect of tropospheric O3. Estimates of present-day

forcing by O3 depend upon currently available information on pre-industrial and current O3

concentrations. Both are limited and, therefore, very uncertain. Other factors, including the

albedo of underlying surface, altitude and co-occurrence of PM can also complicate the

calculation of globally-averaged forcing.
E-33

! Globally-averaged direct forcing by O3. On the basis of the best available information, a

2001 Intergovernmental Panel on Climate Change (IPCC) report offered an estimated value

of 0.35 ± 0.15 Wm!2 for the annual, globally-averaged direct forcing by tropospheric O3. 

Another recent estimate places this value at 0.5 ± 0.2 Wm!2

. 

! Projections of forcing by O3 into the future. A CTM-climate modeling intercomparison

study carried out as part of the third assessment by the IPCC yielded an estimated 0.4 to

0.78 Wm!2

 forcing by O3 by the year 2100. The authors of this study concluded that O3 can

be expected to be an important contributor to climate forcing into the future. 

! Climate forcing by O3 at the regional scale may be its most import impact on climate. 

Satellites have detected high O3 concentrations localized at the regional scale that are

associated with large urban centers and extensive biomass burning. Climate forcing by these

high, regional-scale O3 concentrations have been estimated to be on the order of 1 Wm!2 (a

substantial fraction of the direct, globally-averaged forcing due to well-mixed GHGs,

including CO2). The impact of climate forcing at this level depends upon the particular

characteristics of the region in which it occurs. At present, regional-scale modeling studies

are not available that provide estimates of these effects. 

E.7 MATERIALS DAMAGE

The Chapter 11 discussion of O3 effects on man-made materials mainly summarizes key

information from the 1996 O3 AQCD, given that little new pertinent research information on O3-

related materials damage has been published since then. Key points include the following:

! Ozone and other photochemical oxidants react with many economically important man-made

materials, decreasing their useful life and aesthetic appearance. Materials damaged by O3

include elastomers; textiles and fibers; dyes, pigments, and inks; and paints and other surface

coatings. 
E-34

! Elastomeric compounds (natural rubber and synthetic polymers and copolymers of

butadiene, isoprene, and styrene) are highly susceptible, to even low O3 concentrations. 

These compounds are damaged by O3 breaking molecular chains at the carbon-carbon double

bond and by adding a chain of three oxygen atoms directly across the double bond. This

structure change promotes characteristic cracking of stressed/stretched rubber called

“weathering.” Tensile strain produces cracks on the surface of the rubber that increase in

size and number with increased stress/stretching. The rate of crack growth is dependent on

degree of stress, type of rubber compound, O3 concentration, duration of exposure, O3

velocity, and temperature. After initial cracking, further O3 penetration results in additional

cracking and, eventually, mechanical weakening.

! Ozone can damage textiles and fabrics by mechanisms similar to those associated with

elastomers. Generally, synthetic fibers are less affected by O3 than natural fibers. Overall,

O3 contribution to degradation of textiles and fabrics is not considered significant.

! Ozone fading of textile dyes is a diffusion-controlled process, with the rate of fading being

controlled by diffusion of the dye to the fiber surface. Many textile dyes react with O3. The

rate and severity of the O3 attack is influenced by the chemical nature of the textile fiber. 

! Paints applied to exterior surfaces of buildings and other structures (e.g., bridges), as well as

several artists’ pigments, are also sensitive to fading and oxidation by O3 at concentrations

found in U.S. urban areas. 
1. INTRODUCTION

This is an update revision of the document, “Air Quality Criteria for Ozone and Related 

Photochemical Oxidants,” published by the U.S. Environmental Protection Agency (EPA) in 

1996 (U.S. Environmental Protection Agency, 1996a). That 1996 Ozone Air Quality Criteria 

Document (O3 AQCD) provided scientific bases for Congressionally-mandated periodic review 

by the EPA of the Ozone National Ambient Air Quality Standards (O3 NAAQS), which led to 

promulgation of new O3 NAAQS by EPA in 1997 (Federal Register, 1997). 

The present document critically assesses the latest scientific information relative to 

characterizing health and welfare effects associated with the presence of various concentrations 

of O3 and related oxidants in ambient air. It builds upon the previous 1996 EPA O3 AQCD, 

by focusing on evaluation and integration of scientific information relevant to O3 NAAQS 

criteria development that has become available since that covered by the 1996 criteria review; 

and it will provide scientific bases for the current periodic review of the O3 NAAQS. 

This introductory chapter of the revised O3 AQCD presents: (a) background information 

on legislative requirements, the criteria and NAAQS review process, and the history of O3 

NAAQS reviews (including a chronology of changes in key elements of the O3 standards); 

(b) an overview of the current O3 criteria review process and associated key milestones; and 

(c) an orientation to the general organizational structure and content of the document. 

1.1 LEGAL AND HISTORICAL BACKGROUND 

1.1.1 Legislative Requirements 

Two sections of the U.S. Clean Air Act (CAA) govern establishment, review, and revision 

of National Ambient Air Quality Standards (NAAQS). Section 108 of the CAA (42 U.S.C. 

7408) directs the Administrator of the U.S. Environmental Protection Agency (EPA) to identify 

ambient air pollutants that may be reasonably anticipated to endanger public health or welfare 

and to issue air quality criteria for them. The air quality criteria are to reflect the latest scientific 

information useful in indicating the kind and extent of all identifiable effects on public health or 

welfare that may be expected from the presence of a given pollutant in ambient air. 

1-1
Section 109(a) of the CAA (42 U.S.C. 7409) directs the Administrator of EPA to propose 

and promulgate primary and secondary NAAQS for pollutants identified under Section 108. 

Section 109(b)(1) defines a primary standard as one that, in the judgment of the Administrator, is 

requisite to protect the public health (see inset below) based on the criteria and allowing for an 

adequate margin of safety. The secondary standard, as defined in Section 109(b)(2), must 

specify a level of air quality that, in the judgment of the Administrator, is requisite to protect the 

public welfare (see inset below) from any known or anticipated adverse effects associated with 

the presence of the pollutant in ambient air, based on the criteria. 

EXAMPLES OF EXAMPLES OF 

PUBLIC HEALTH EFFECTS PUBLIC WELFARE EFFECTS 

# Effects on the health of the general population, # Effects on personal comfort and well-being 

or identifiable groups within the population, # Effects on economic values 

who are exposed to pollutants in ambient air # Deterioration of property 

# Effects on mortality # Hazards to transportation 

# Effects on morbidity # Effects on the environment, including: 

# Effects on other health conditions including 

indicators of: C animals C vegetation 

C climate C visibility 

C pre-morbid processes, C crops C water 

C risk factors, and C materials C weather 

C disease C soils C wildlife 

Section 109(d) of the CAA (42 U.S.C. 7409) requires periodic review and, if appropriate, 

revision of existing criteria and standards. If, in the Administrator's judgment, the Agency's 

review and revision of criteria make appropriate the proposal of new or revised standards, such 

standards are to be revised and promulgated in accordance with CAA Section 109(b). Or, the 

Administrator may find that revision of the standards is not appropriate and conclude the review 

by leaving existing standards unchanged. Section 109(d)(2) of the CAA also requires that an 

independent scientific review committee be established to advise the EPA Administrator on 

NAAQS matters, including the scientific soundness of criteria (scientific bases) supporting 

NAAQS decisions. This role is fulfilled by the Clean Air Scientific Advisory Committee 

(CASAC), which is administratively supported by EPA’s Science Advisory Board (SAB). 

1-2
1.1.2 Criteria and NAAQS Review Process 

Periodic reviews by EPA of criteria and NAAQS for a given criteria air pollutant progress 

through a number of steps, beginning with preparation of an air quality criteria document 

(AQCD) by EPA’s National Center for Environmental Assessment Division in Research 

Triangle Park, NC (NCEA-RTP). The AQCD provides a critical assessment of the latest 

available scientific information upon which the NAAQS are to be based. Drawing upon the 

AQCD, staff of EPA’s Office of Air Quality Planning and Standards (OAQPS) prepare a Staff 

Paper that evaluates policy implications of the key studies and scientific information contained 

in the AQCD and presents EPA staff conclusions and recommendations for standard-setting 

options for the EPA Administrator to consider. The Staff Paper is intended to help “bridge the 

gap” between the scientific assessment contained in the AQCD and the judgments required of 

the Administrator in determining whether it is appropriate to retain or to revise the NAAQS. 

Iterative drafts of the AQCD and the Staff Paper (as well as other analyses, such as exposure 

and/or risk assessments, supporting the Staff Paper) are made available for public comment and 

CASAC review. Final versions of the AQCD and Staff Paper incorporate changes made in 

response to CASAC and public review. Based on the information in these documents, the EPA 

Administrator proposes decisions on whether to retain or revise the NAAQS, taking into account 

public comments and CASAC advice and recommendations. The Administrator’s proposed 

decisions are published in the Federal Register, with a preamble that presents the rationale for 

the decisions and solicits public comment. After considering comments received on the 

proposed decisions, the Administrator then makes final decisions on retaining or revising the 

NAAQS, which are promulgated in a Federal Register notice that addresses significant 

comments received on the proposal. 

NAAQS decisions involve consideration of the four basic elements of a standard: 

indicator, averaging time, form, and level. The indicator defines the pollutant to be measured in 

the ambient air for the purpose of determining compliance with the standard. The averaging 

time defines the time period over which air quality measurements are to be obtained and 

averaged, considering evidence of effects associated with various time periods of exposure. 

The form of a standard defines the air quality statistic that is to be compared to the level of the 

standard (i.e., an ambient concentration of the indicator pollutant) in determining whether an 

area attains the standard. The form of the standard specifies the air quality measurements that 

1-3
are to be used for compliance purposes (e.g., the 98th percentile of an annual distribution of 

daily concentrations; the annual arithmetic average), the monitors from which the measurements 

are to be obtained (e.g., one or more population-oriented monitors in an area), and whether the 

statistic is to be averaged across multiple years. These basic elements of a standard are the 

primary focus of the staff conclusions and recommendations in the Staff Paper and in the 

subsequent rulemaking, building upon the policy-relevant scientific information assessed in the 

AQCD and on the policy analyses contained in the Staff Paper. These four elements taken 

together determine the degree of public health and welfare protection afforded by the NAAQS. 

1.1.3 Regulatory Chronology1 

On April 30, 1971, primary and secondary NAAQS for photochemical oxidants were 

promulgated by EPA under Section 109 of the CAA (36 FR 8186). These NAAQS were set at 

an hourly average of 0.08 ppm total photochemical oxidants, not to be exceeded more than 1 h 

per year. On April 20, 1977, the EPA announced (42 FR 20493) the first review and updating of 

the 1970 Air Quality Criteria Document for Photochemical Oxidants in accordance with Section 

109(d) of the CAA. In preparing that criteria document, EPA made two external review drafts of 

the document available for public comment, and these drafts were peer reviewed by the 

Subcommittee on Scientific Criteria for Photochemical Oxidants of EPA’s Science Advisory 

Board (SAB). A final revised AQCD for Ozone and Other Photochemical Oxidants was then 

published on June 22, 1978. 

Based on the revised 1978 AQCD and taking into account the advice and recommendations 

of the SAB Subcommittee and public comments, the EPA announced (44 FR 8202) a final 

decision to revise the NAAQS for photochemical oxidants on February 8, 1979. That final 

rulemaking revised the primary standard from 0.08 ppm to 0.12 ppm, set the secondary standard 

to be the same as the primary standard, changed the chemical designation of the standards from 

photochemical oxidants to O3, and revised the definition of the point at which the standard is 

attained as indicated in Table 1-1. 

1

This following text is excerpted and adapted from the “Proposed Decision on the National Ambient 

Air Quality Standards for Ozone,” 57 FR 35542, 35542-35557 (August, 10, 1992) and the “National Ambient Air 

Quality Standards for Ozone; Final Rule,” 62 FR 38856, 83356-38896 (July 18, 1997). 

1-4 
Table 1-1. National Ambient Air Quality Standards (NAAQS) for Ozone 

Date of Promulgation Primary and Secondary NAAQS Averaging Time 

February 8, 1979 0.12 ppma

 (235 μg/m3) 1 hb 

July 18, 1997 0.08 ppma (157 μg/m3) 8 hc 

a

1 ppm = 1962 μg/m3, 1 μg/m3 = 5.097 × 10!4 ppm @ 25 °C, 760 mm Hg. 

b

The standard is attained when the expected number of days per calendar year with a maximum hourly average

 concentration above 235 μg/m3

 (0.12 ppm) is equal to or less than one.

c

Based on the 3-year average of the annual fourth-highest daily maximum 8-h average concentration measured

 at each monitor within an area.

Source: Federal Register (1979, 1997). 

On March 17, 1982, in response to requirements of Section 109(d) of the CAA, the EPA 

announced (47 FR 11561) that it planned to revise the existing 1978 AQCD for Ozone and Other 

Photochemical Oxidants; and, on August 22, 1983, it announced (48 FR 38009) that review of 

the primary and secondary NAAQS for O3 had been initiated. The EPA provided a number of 

opportunities for expert review and public comment on revised chapters of the AQCD, including 

two public peer-review workshops in December 1982 and November 1983. Comments made at 

both workshops were considered by EPA in preparing the First External Review Draft that was 

made available (49 FR 29845) on July 24, 1984, for public review. On February 13, 1985 

(50 FR 6049) and then on April 2, 1986 (51 FR 11339), EPA announced two public CASAC 

meetings, which were held on March 4-6, 1985 and April 21-22, 1986, respectively. At these 

meetings, the CASAC reviewed external review drafts of the revised AQCD for O3 and Other 

Photochemical Oxidants. After these two reviews, CASAC’s consensus views were summarized 

by the CASAC Chair in an October 1986 letter to the EPA Administrator, which stated that the 

document “represents a scientifically balanced and defensible summary of the extensive 

scientific literature.” Taking into account public and CASAC comments on the two external 

review drafts, revisions were made by EPA and the final document was released by EPA in 

August 1986. 

The first draft of the Staff Paper “Review of the National Ambient Air Quality Standards 

for Ozone: Assessment of Scientific and Technical Information” drew upon key findings and 

conclusions from the AQCD and was reviewed by CASAC at the April 21-22, 1986 public 

1-5
meeting. At that meeting, the CASAC recommended that new information on prolonged O3 

exposure effects be considered in a second draft of the Staff Paper. The CASAC reviewed the 

resulting second draft and also heard a presentation of new and emerging information on the 

health and welfare effects of O3 at a December 14-15, 1987 public review meeting. The CASAC 

concluded that sufficient new information existed to recommend incorporation of relevant 

new data into a supplement to the 1986 AQCD (O3 Supplement) and into a third draft of the 

Staff Paper. 

A draft O3 Supplement, entitled “Summary of Selected New Information on Effects of 

Ozone on Health and Vegetation: Draft Supplement to Air Quality Criteria for Ozone and Other 

Photochemical Oxidants,” and the revised Staff Paper were made available to CASAC and to 

the public in November 1988. The O3 Supplement assessed selected literature concerning 

exposure- and concentration-response relationships observed for health effects in humans and 

experimental animals and for vegetation effects that appeared in papers published or in-press 

from 1986 through early 1989. On December 14-15, 1988, the CASAC held a public meeting to 

review these documents and then sent the EPA Administrator a letter (dated May 1, 1989), 

which stated that the draft O3 Supplement, the 1986 AQCD, and the draft Staff Paper “provide 

an adequate scientific basis for the EPA to retain or revise the primary and secondary standards 

of ozone.” The CASAC concluded (a) that it would be some time before sufficient new 

information on the health effects of multihour and chronic exposure to O3 would be published in 

scientific journals to receive full peer review and, thus, be suitable for inclusion in a criteria 

document and (b) that such information could be considered in the next review of the O3 

NAAQS. A final version of the O3 Supplement was published in 1992 (U.S. Environmental 

Protection Agency, 1992). 

On October 22, 1991, the American Lung Association and other plaintiffs filed suit to 

compel the Agency to complete the review of the criteria and standards for O3 in accordance 

with the CAA. The U.S. District Court for the Eastern District of New York subsequently issued 

an order requiring the EPA to announce its proposed decision on whether to revise the standards 

for O3 by August 1, 1992 and to announce its final decision by March 1, 1993. 

The proposed decision on O3, which appeared in the Federal Register on August 10, 1992 

(57 FR 35542), indicated that revision of the existing 1-h O3 NAAQS was not appropriate at that 

time. A public hearing on this decision was held in Washington, DC on September 1, 1992; and 

1-6
public comments were received through October 9, 1992. The final decision not to revise the 

1-h NAAQS was published in the Federal Register on March 9, 1993 (58 FR 13008). However, 

that decision did not take into consideration a number of more recent studies on the health 

and welfare effects of O3 that had been published since the last of the literature assessed in 

the O3 Supplement (i.e., studies available through 1985 and into early 1986). 

The Agency initiated consideration of such studies as part of the next congressionallymandated periodic review of O3 criteria and NAAQS. The new studies were assessed in revised 

draft O3 AQCD chapters that were peer reviewed in July and September 1993 workshops, 

followed by public release of the O3 AQCD First External Review Draft in February 1994 and 

CASAC review on July 20-21, 1994. Further drafts of the O3 AQCD, revised in response to 

public comments and CASAC review, were reviewed by CASAC on March 21-25, 1995, and at 

a final CASAC review meeting on September 19-20, 1995. The scientific soundness of the 

revised O3 AQCD was recognized by CASAC in a November 28, 1995 letter to the EPA 

Administrator; and the final O3 AQCD was published in July 1996. 

The first draft of the associated Staff Paper, “Review of the National Ambient Air Quality 

Standards for Ozone: Assessment of Scientific and Technical Information,” was also reviewed 

by CASAC at the March 21-22, 1995 public meeting. CASAC also reviewed subsequent drafts 

of the Staff Paper at public meetings on September 19-20, 1995 and March 21, 1996, with 

completion of CASAC review of the primary and secondary standard portions of the draft 

Staff Paper being communicated in letters to the EPA Administrator dated November 30, 1995 

and April 4, 1996, respectively. The final O3 Staff Paper was published in June 1996 (U.S. 

Environmental Protection Agency, 1996b). 

On December 13, 1996, EPA published its proposed decision to revise the O3 NAAQS 

(61 FR 65716). Extensive opportunities for public comment on the proposed decision, including 

several public hearings and two national satellite telecasts, were then provided by EPA; and 

EPA’s final decision to promulgate a new 8-h O3 NAAQS (see Table 1-1) was published on July 

18, 1997 (62 FR 38856). 

Following promulgation of the new standards, numerous petitions for review of the 

standards were filed in the U.S. Court of Appeals for the District of Columbia Circuit (D.C. 

Circuit)2

. On May 14, 1999, the Court remanded the O3 NAAQS to EPA, finding that 

2

American Trucking Associations v. EPA, No. 97-1441. 

1-7 
Section 109 of the CAA, as interpreted by EPA, effected an unconstitutional delegation of 

legislative authority3. In addition, the Court directed that, in responding to the remand, EPA 

should consider the potential beneficial health effects of O3 pollution in shielding the public from 

the effects of solar ultraviolet (UV) radiation. On January 27, 2000, EPA petitioned the U.S. 

Supreme Court for certiorari on the constitutional issue (and two other issues), but did not 

request review of the D.C. Circuit ruling regarding the potential beneficial health effects of O3. 

On February 27, 2001, the U.S. Supreme Court unanimously reversed the judgment of the D.C. 

Circuit on the constitutional issue (holding that section 109 of the CAA does not delegate 

legislative power to the EPA in contravention of the Constitution) and remanded the case to the 

D.C. Circuit to consider challenges to the O3 NAAQS that had not been addressed by that Court's 

earlier decisions4

. On March 26, 2002, the D.C. Circuit issued its final decision, finding that the 

1997 O3 NAAQS were “neither arbitrary nor capricious,” and denied the remaining petitions 

for review5

. 

On November 14, 2001, EPA proposed to respond to the Court’s remand to consider the 

potential beneficial health effects of O3 pollution in shielding the public from the effects of solar 

UV radiation by leaving the 1997 8-h NAAQS unchanged. Following a review of information in 

the record and the substantive comments received on the proposed response, EPA issued a final 

response to the remand, reaffirming the 8 h O3 NAAQS (68 FR 614, January 6, 2003). 

1.2 CURRENT OZONE CRITERIA AND NAAQS REVIEW 

1.2.1 Key Milestones and Procedures for Document Preparation 

It is important to note at the outset that development of the present O3 AQCD included 

substantial external expert review and opportunities for public input through (a) public 

workshops involving the general scientific community, (b) iterative reviews of successive 

AQCD drafts by CASAC, and (c) comments from the public on successive drafts. Extensive 

external inputs received through such reviews help to ensure that the current periodic review of 

3 American Trucking Associations v. EPA, 175 F.3d 1027 (D.C. Cir., 1999). 

4

Whitman v. American Trucking Associations, 531 U.S. 457 (2001). 

5

American Trucking Associations v. EPA, 283 F.3d 355, (D.C. Cir. 2002). 

1-8 
the O3 standards is based on critical assessment of the latest available pertinent science as 

presented in this document and drawn upon in the associated Ozone Staff Paper. 

The procedures for developing this revised O3 AQCD build on experience derived from the 

other recent criteria document preparation efforts, with key milestones for development of 

this O3 AQCD being listed in Table 1-2. Briefly, respective responsibilities for production of the 

document and key milestones are as follows. An NCEA-RTP Ozone Team created and 

implemented a project plan for developing the O3 AQCD, taking into account input from 

individuals in other EPA program and policy offices identified as part of the EPA Ozone Work 

Group. The resulting plan, i.e., the “Project Work Plan for Revised Air Criteria for Ozone and 

Related Photochemical Oxidants” (November 2002), was discussed with CASAC in January 

2003. Under the processes established in Sections 108 and 109 of the CAA, the EPA officially 

initiated the current criteria and NAAQS review by announcing the commencement of the 

review in the Federal Register (65 FR 57810, September, 2000) with a call for information. That 

Federal Register notice included (1) a request asking for recently available research information 

on O3 that may not yet have been published and (2) a request for individuals with the appropriate 

type and level of expertise to contribute to the writing of O3 AQCD materials to identify 

themselves. The specific authors of chapters or sections of the proposed document included both 

EPA and non-EPA scientific experts, who were selected on the basis of their expertise on the 

subject areas and their familiarity with the relevant literature. The project team defined critical 

issues and topics to be addressed by the authors and provided direction in order to focus on 

evaluation of those studies most clearly identified as important for standard setting. An ongoing 

literature search that was underway prior to initiation of work on this document continued 

throughout its preparation to identify pertinent O3 literature published since early 1996. 

As with other NAAQS reviews, critical assessment of relevant scientific information is 

presented in this updated O3 AQCD. The main focus of this document is the evaluation and 

interpretation of pertinent atmospheric science information, air quality data, human exposure 

information, and health and welfare effects information newly published since that assessed in 

the 1996 O3 AQCD. Draft versions of AQCD chapter materials were evaluated via expert peerconsultation workshop discussions (see Table 1-2) that focused on the selection of pertinent 

studies to be included in the chapters, the potential need for additional information to be added to 

the chapters, and the quality of the characterization and interpretation of the literature. The 

1-9
Table 1-2. Key Milestones for Development of Revised Ozone Air Quality 

Criteria Document (O3 AQCD) 

Major Milestones 

 1. Literature Search 

 2. Federal Register Call for Information 

 3. Draft Project Plan Available for Public Comment 

 4. Revised Draft Project Plan Released for CASAC Consultation 

 5. CASAC Consultation on Draft Project Work Plan 

 6. Peer-Consultation Workshop on Draft Ecological Effects Materials 

 7. Peer-Consultation Workshops on Draft Atmospheric

Science/Exposure and Dosimetry/Health Chapters

 8. First External Review Draft of O3 AQCD 

 9. Public Comment Period (90 days) 

10. CASAC Public Review Meeting (First External Review Draft) 

11. Second External Review Draft of O3 AQCD 

12. Public Comment Period (90 days) 

13. CASAC Public Review Meeting (Second External Review Draft) 

14. Final O3 AQCD 

Dates

Ongoing

September 2000

Dec 2001 - March 2002

December 2002

January 2003

April 2003

July 2004

January 2005

Feb - April 2005 

May 4-5, 2005 

August 2005 

Sept - Nov 2005 

December 6-8, 2005 

February 28, 2006 

authors of the draft chapters then revised them on the basis of the workshop and/or other expert 

review comments6

. These and other integrative materials were then incorporated into the First 

External Review Draft (January 2005) of this O3 AQCD, which was made available for public 

comment and CASAC review, as indicated in Table 1-2. 

Following review of the First External Review Draft at a May 4-5, 2005 CASAC meeting, 

EPA incorporated revisions into the draft O3 AQCD in response to comments from CASAC and 

the public and made a Second External Review Draft (August, 2005) available for further public 

comment and CASAC review as shown in Table 1-2. More specifically, the Second External 

Review Draft underwent public comment during September-November, 2005, and was reviewed 

by CASAC at a December 6-8, 2005 public meeting. This final O3 AQCD, completed by 

6

It should be noted that materials contributed by non-EPA authors were, at times, modified by EPA Ozone 

Team staff in response to internal and/or external review comments and that EPA is responsible for the ultimate 

content of this O3 AQCD. 

1-10 
February 28, 2006, incorporates revisions made in response to public comments and CASAC 

reviews of the earlier draft AQCD materials. An electronic version of this document can be 

accessed via an EPA website at: www.epa.gov/ncea. 

The EPA’s Office of Air Quality Planning and Standards (OAQPS) staff is also preparing 

further draft O3 Staff Paper materials which draw upon key information contained in this final O3 

AQCD. After review of that draft O3 Staff Paper by the public and by CASAC, EPA will take 

public and CASAC comments into account in producing a Final Ozone Staff Paper. That Staff 

Paper, in final form, will present options for consideration by the Administrator of EPA 

regarding whether to retain or, if appropriate, to revise the O3 NAAQS. 

1.3 ORGANIZATIONAL STRUCTURE OF THE DOCUMENT 

1.3.1 General Document Format 

The general format used in preparing this O3 AQCD is to open each new section for the 

updated document with concise summarization of key findings and conclusions from the 

previous 1996 O3 AQCD (U.S. Environmental Protection Agency, 1996a). After presentation of 

such background information, the remainder of each section typically provides an updated 

discussion of newer literature and resulting key conclusions. In some cases where no new 

information is available, the summary of key findings and conclusions from the previous criteria 

document must suffice as the basis for current key conclusions. Increased emphasis is placed in 

the main chapters of this revised O3 AQCD on interpretative evaluation and integration of 

evidence pertaining to a given topic than has been typical of previous EPA air quality criteria 

documents, with more detailed descriptions of individual studies being provided in a series of 

accompanying annexes. 

A list of references published since completion of the 1996 criteria document was made 

available to the authors. The references were selected from information data base searches 

conducted by EPA. Additional references were added to the list (e.g., missed or recently 

published papers or “in press” publications) as work proceeded in creating draft document 

materials. As an aid in selecting pertinent new literature, the authors were also provided with a 

summary of issues that needed to be addressed in this revised O3 AQCD. These issues were 

identified by NCEA-RTP Ozone Team members, by the EPA Ozone Work Group, and by 

1-11
authors and reviewers of draft O3 AQCD materials, and they were further expanded, as 

appropriate, based on public discussions, workshops, or other comments received by EPA in the 

course of development of this document. 

1.3.2 Organization and Content of the Document 

This revised AQCD for Ozone and Related Photochemical Oxidants critically assesses 

scientific information on the health and welfare effects associated with exposure to the 

concentrations of these pollutants in ambient air. The document does not provide a detailed 

literature review; but, rather, discusses cited references that reflect the current state of knowledge 

on the most relevant issues pertinent to the derivation of NAAQS for O3 and/or related 

photochemical oxidants. Although emphasis is placed on discussion of health and welfare 

effects information, other scientific data are presented and evaluated in order to provide a better 

understanding of the nature, sources, distribution, measurement, and concentrations of O3 and 

related photochemical oxidants in ambient air, as well as the measurement of population 

exposure to these pollutants. 

The main focus of the scientific information discussed in the text comes from literature 

published since completion of the 1996 O3 AQCD (U.S. Environmental Protection Agency, 

1996a). Emphasis is placed on studies conducted at or near O3 concentrations found in ambient 

air. Other studies are included if they contain unique data, such as the documentation of a 

previously unreported effect or of a mechanism for an observed effect; or if they were multipleconcentration studies designed to elucidate exposure-response relationships. Generally, this is 

not an issue for human clinical or epidemiology studies. However, for animal toxicology 

studies, consideration is given mainly to those studies conducted at less than 1 ppm O3. 

Key information from studies assessed in the previous O3 AQCD and whose data impacted the 

derivation of the current NAAQS are briefly summarized in the text, along with specific citations 

to the previous document. Prior studies are also discussed if they (1) are open to reinterpretation 

in light of newer data, or (2) are potentially useful in deriving revised standards for O3. 

Generally, only information that has undergone scientific peer review and has been published 

(or accepted for publication) through December 2004 is included in this draft document. A few 

particularly pertinent and important new studies published or accepted for publication beyond 

the end of 2004 are also considered. 

1-12
This revised O3 AQCD consists of three volumes. The first volume includes an Executive 

Summary and Conclusions, as well as Chapters 1 through 11 of the document. This introductory 

chapter (Chapter 1) presents background information on the purpose of the document, legislative 

requirements, and the history of past O3 NAAQS regulatory actions, as well as an overview of 

the organization and content of the document. Chapter 2 provides information on the physics 

and chemistry of O3 and related photochemical oxidants in the atmosphere. Chapter 3 covers 

tropospheric O3 environmental concentrations, patterns, and exposures. The accompanying 

annexes to each of these background chapters are found in Volume II. 

Health information pertinent to derivation of the primary O3 NAAQS is then mainly 

covered in the next several chapters (Chapters 4 through 8). Chapter 4 discusses O3 dosimetry 

aspects; and Chapters 5, 6, and 7 discuss animal toxicological studies, controlled-exposure 

studies of human health effects, and epidemiologic studies of ambient air exposure effects on 

human populations, respectively. Chapter 8 then provides an integrative and interpretive 

evaluation of key information relevant to O3 exposure and health risks of most pertinence to the 

review of primary O3 NAAQS. The annexes to these health-related chapters are found in 

Volume II. 

The remaining three chapters of the document assess welfare effects information pertinent 

to the review of secondary O3 NAAQS. More specifically, Chapter 9 deals with ecological and 

other environmental effects of O3 and related photochemical oxidants. Chapter 10 assesses 

tropospheric O3 involvement in climate change processes, including impacts on solar UV flux in 

Earth’s lower atmosphere. Lastly, Chapter 11 discusses O3 effects on man-made materials as a 

third type of welfare effect of potential concern. Annex materials related to welfare effects 

(especially vegetation/ecological effects) are contained in Volume III. 

1-13
REFERENCES 

Federal Register. (1971) National primary and secondary ambient air quality standards. F. R. (April 30) 

36: 8186-8201.

Federal Register. (1977) Review of the photochemical oxidant and hydrocarbon air quality standards. F. R. 

(April 20) 42: 20493-20494. 

Federal Register. (1979) National primary and secondary ambient air quality standards: revisions to the national 

ambient air quality standards for photochemical oxidants. F. R. (February 8) 44: 8202-8237. 

Federal Register. (1982) Air quality criteria document for ozone and other photochemical oxidants. F. R. (March 17) 

47: 11561.

Federal Register. (1983) Review of the national ambient air quality standards for ozone. F. R. (August 22) 

48: 38009.

Federal Register. (1984) Draft air quality criteria document for ozone and other photochemical oxidants. F. R. 

(July 24) 49: 29845. 

Federal Register. (1985) Science Advisory Board; Clean Air Scientific Advisory Committee; open meeting. F. R. 

(February 13) 50: 6049. 

Federal Register. (1986) Science Advisory Board; Clean Air Scientific Advisory Committee; open meeting. F. R. 

(April 2) 51: 11339. 

Federal Register. (1992) National ambient air quality standards for ozone; proposed decision. F. R. (August 10) 

57: 35542-35557.

Federal Register. (1993) National ambient air quality standards for ozone - final decision. F. R. (March 9) 

58: 13008-13019.

Federal Register. (1996) National ambient air quality standards for ozone: proposed decision. F. R. (December 13) 

61: 65,716-65,750.

Federal Register. (1997) National ambient air quality standards for ozone; final rule. F. R. (July 18) 

62: 38856-38896.

Federal Register. (2000) Air Quality Criteria for Ozone and Related Photochemical Oxidants; notice; call for 

information. F. R. (September 26) 65: 57810. 

Federal Register. (2003) National ambient air quality standards for ozone: final response to remand; final rule. F. R. 

(January 6) 68: 614-645. 

U.S. Code. (2003a) Clean Air Act, §108, air quality criteria and control techniques.. U. S. C. 42: §7408. 

U.S. Code. (2003b) Clean Air Act, §109, national ambient air quality standards. U. S. C. 42: §7409. 

U.S. Court of Appeals for the District of Columbia. (1999a) American Trucking Associations, Inc. v. U.S. 

Environmental Protection Agency. 195 F.3d 4 (D.C. Cir. 1999). 

U.S. Court of Appeals for the District of Columbia. (1999b) American Trucking Associations, Inc. v. U.S. 

Environmental Protection Agency. 175 F.3d 1027 (D.C. Cir. 1999). 

U.S. Court of Appeals for the District of Columbia. (2002) American Trucking Associations, Inc. v. U.S. 

Environmental Protection Agency. 283 F.3d 355, 378-79 (D.C. Cir. 2002). 

U.S. Environmental Protection Agency. (1992) Summary of selected new information on effects of ozone on health 

and vegetation: supplement to 1986 air quality criteria for ozone and other photochemical oxidants. Research 

Triangle Park, NC: Office of Health and Environmental Assessment, Environmental Criteria and Assessment 

Office; report no. EPA/600/8-88/105F. Available from: NTIS, Springfield, VA; PB92-235670. 

U.S. Environmental Protection Agency. (1996a) Air quality criteria for ozone and related photochemical oxidants. 

Research Triangle Park, NC: Office of Research and Development; report nos. EPA/600/AP-93/004aF-cF. 3v. 

Available from: NTIS, Springfield, VA; PB96-185582, PB96-185590, and PB96-185608. Available: 

http://cfpub2.epa.gov/ncea/. 

U.S. Environmental Protection Agency. (1996b) Review of national ambient air quality standards for ozone: 

assessment of scientific and technical information. OAQPS staff paper. Research Triangle Park, NC: 

Office of Air Quality Planning and Standards; report no. EPA/452/R-96/007. Available from: NTIS, 

Springfield, VA; PB96-203435. Available: http://www.epa.gov/ttn/naaqs/standards/ozone/s_o3_pr_sp.html 

(29 September 2005). 

U.S. Supreme Court. (2001) Whitman v. American Trucking Association. 531 U.S. 457 (nos. 99-1257 and 99-1426). 

1-14
2. PHYSICS AND CHEMISTRY OF OZONE

IN THE ATMOSPHERE

2.1 INTRODUCTION 

Ozone (O3) and other oxidants, such as peroxacyl nitrates and hydrogen peroxide (H2O2) 

form in polluted areas by atmospheric reactions involving two main classes of precursor 

pollutants, volatile organic compounds (VOCs) and nitrogen oxides (NOx). Carbon monoxide 

(CO) is also important for O3 formation in polluted areas. Ozone is thus a secondary pollutant. 

The formation of O3, other oxidants and oxidation products from these precursors is a complex, 

nonlinear function of many factors: the intensity and spectral distribution of sunlight; 

atmospheric mixing and processing on cloud and aerosol particles; the concentrations of the 

precursors in ambient air; and the rates of chemical reactions of the precursors. Information 

contained in this chapter and in greater detail in Annex AX2 describes these processes, 

numerical models that incorporate these processes to calculate O3 concentrations, and techniques 

for measuring concentrations of ambient oxidants. 

The atmosphere can be divided into several distinct vertical layers, based primarily on the 

major mechanisms by which they are heated and cooled. The lowest major layer is the 

troposphere, which extends from the earth’s surface to about 8 km above polar regions and to 

about 16 km above tropical regions. The planetary boundary layer (PBL) is the lower sublayer 

of the troposphere, extending from the surface to about 1 or 2 km and is most strongly affected 

by surface conditions. The stratosphere extends from the tropopause, or the top of the 

troposphere, to about 50 km in altitude (Annex AX2.2.1). The emphasis in this chapter is placed 

on chemical and physical processes occurring in the troposphere, in particular in the PBL. The 

processes responsible for producing summertime O3 episodes are fairly well understood, as 

discussed in the previous Air Quality Criteria Document for Ozone and Related Photochemical 

Oxidants or 1996 O3 AQCD (U.S. Environmental Protection Agency, 1996). This chapter 

mainly considers topics for which there is substantial new information and on topics that form 

the basis for discussions in later chapters. 

2-1
2.2 CHEMICAL PROCESSES INVOLVED IN OZONE FORMATION 

AND DESTRUCTION 

Ozone occurs not only in polluted urban atmospheres but throughout the troposphere, even 

in remote areas of the globe. The same basic processes, involving sunlight-driven reactions 

of NOx and VOCs contribute to O3 formation throughout the troposphere. These processes also 

lead to the formation of other photochemical products, such as peroxyacetyl nitrate (PAN), nitric 

acid (HNO3), and sulfuric acid (H2SO4), and to other compounds, such as formaldehyde (HCHO) 

and other carbonyl compounds, such as aldehydes and ketones. 

The photochemical formation of O3 in the troposphere proceeds through the oxidation of 

nitric oxide (NO) to nitrogen dioxide (NO2) by organic (RO2) or hydro-peroxy (HO2) radicals. 

The photolysis of NO2 yields nitric oxide (NO) and a ground-state oxygen atom, O(3P), which 

then reacts with molecular oxygen to form O3. Free radicals oxidizing NO to NO2 are formed 

during the oxidation of VOCs (Annex AX2.2.2). 

The term VOC refers to all carbon-containing gas-phase compounds in the atmosphere, 

both biogenic and anthropogenic in origin, excluding CO and CO2. Classes of organic 

compounds important for the photochemical formation of O3 include alkanes, alkenes, aromatic 

hydrocarbons, carbonyl compounds (e.g., aldehydes and ketones), alcohols, organic peroxides, 

and halogenated organic compounds (e.g., alkyl halides). This array of compounds encompasses 

a wide range of chemical properties and lifetimes: isoprene has an atmospheric lifetime of 

approximately an hour, whereas methane has an atmospheric lifetime of about a decade. 

In urban areas, compounds representing all classes of VOCs and CO are important for O3 

formation. In nonurban vegetated areas, biogenic VOCs emitted from vegetation tend to be the 

most important. In the remote troposphere, CH4 and CO are the main carbon-containing 

precursors to O3 formation. CO also can play an important role in O3 formation in urban areas. 

The oxidation of VOCs is initiated mainly by reaction with hydroxyl (OH) radicals. The primary 

source of OH radicals in the atmosphere is the reaction of electronically excited O atoms, O(1

D), 

with water vapor. O(1

D) is produced by the photolysis of O3 in the Hartley bands. In polluted 

areas, the photolysis of aldehydes (e.g., HCHO), nitrous acid (HONO) and hydrogen 

peroxide (H2O2) can also be significant sources of OH or HO2 radicals that can rapidly be 

converted to OH (Eisele et al., 1997). Ozone can oxidize alkenes; and, at night, when they are 

most abundant, NO3 radicals also oxidize alkenes. In coastal environments and other selected 

2-2
environments, atomic Cl and Br radicals can also initiate the oxidation of VOCs (Annex 

AX2.2.3). 

There are a large number of oxidized nitrogen containing compounds in the atmosphere 

including NO, NO2, NO3, HNO2, HNO3, N2O5, HNO4, PAN and its homologues, other organic 

nitrates and particulate nitrate. Collectively these species are referred to as NOy. Oxidized 

nitrogen compounds are emitted to the atmosphere mainly as NO which rapidly interconverts 

with NO2 and so NO and NO2 are often “lumped” together into their own group or family, 

or NOx. NOx can be oxidized to reservoir and termination species (PAN and its homologues, 

organic nitrates, HNO3, HNO4 and particulate nitrate). These reservoir and termination species 

are referred to as NOz. The major reactions involving interconversions of oxidized nitrogen 

species are discussed in Annex AX2.2.4. 

The photochemical cycles by which the oxidation of hydrocarbons leads to O3 production 

are best understood by considering the oxidation of methane, structurally the simplest VOC. 

The CH4 oxidation cycle serves as a model for the chemistry of the relatively clean or unpolluted 

troposphere (although this is a simplification because vegetation releases large quantities of 

complex VOCs, such as isoprene, into the atmosphere). In the polluted atmosphere, the 

underlying chemical principles are the same, as discussed in Annex AX2.2.5. The conversion of 

NO to NO2 occurring with the oxidation of VOCs is accompanied by the production of O3 and 

the efficient regeneration of the OH radical, which in turn can react with other VOCs. 

A schematic overview showing the major processes involved in O3 production and loss in the 

troposphere and stratosphere is given in Figure 2-1. 

The oxidation of alkanes and alkenes in the atmosphere has been treated in depth in 

1996 O3 AQCD and is updated in Annexes AX2.2.6 and AX2.2.7. In contrast to simple 

hydrocarbons containing one or two carbon atoms, detailed kinetic information about the gas 

phase oxidation pathways of many anthropogenic hydrocarbons (e.g., aromatic compounds such 

as benzene and toluene), biogenic hydrocarbons (e.g., isoprene, the monoterpenes), and their 

intermediate oxidation products (e.g., epoxides, nitrates, and carbonyl compounds) is lacking. 

Reaction with OH radicals represents the major loss process for alkanes. Reaction with chlorine 

atoms is an additional sink for alkanes. Stable products of alkane photooxidation are known to 

include carbonyl compounds, alkyl nitrates, and d-hydroxycarbonyls. Major uncertainties in the 

atmospheric chemistry of the alkanes concern the chemistry of alkyl nitrate formation; these 

2-3
Figure 2-1. Schematic overview of O3 photochemistry in the stratosphere 

and troposphere. 

uncertainties affect the amount of NO-to-NO2 conversion occurring and, hence, the amounts 

of O3 formed during photochemical degradation of the alkanes. 

The reaction of OH radicals with aldehydes produced during the oxidation of alkanes 

forms acyl (R'CO) radicals, and acyl peroxy radicals (R'C(O)!O2) are formed by the further 

addition of O2. As an example, the oxidation of ethane (C2H5!H) yields acetaldehyde 

(CH3!CHO). The reaction of CH3!CHO with OH radicals yields acetyl radicals (CH3!CO). 

The acetyl radicals will then participate with O2 in a termolecular recombination reaction to form 

acetyl peroxy radicals, which can then react with NO to form CH3 + CO2 or they can react with 

2-4
NO2 to form PAN. PAN acts as a temporary reservoir for NO2. Upon the thermal 

decomposition of PAN, either locally or elsewhere, NO2 is released to participate in the O3 

formation process again. 

Alkenes react in ambient air with OH, NO3, and Cl radicals and with O3. All of these 

reactions are important atmospheric transformation processes, and all proceed by initial addition 

to the >C=C< bonds. Products of alkene photooxidation include carbonyl compounds, 

hydroxynitrates and nitratocarbonyls, and decomposition products from the energy-rich 

biradicals formed in alkene-O3 reactions. Major uncertainties in the atmospheric chemistry of 

the alkenes concern the products and mechanisms of their reactions with O3, especially the yields 

of free radicals that participate in O3 formation. Examples of oxidation mechanisms of complex 

alkanes and alkenes can be found in comprehensive texts such as Seinfeld and Pandis (1998). 

The oxidation of aromatic hydrocarbons constitutes an important component of the 

chemistry of O3 formation in urban atmospheres (Annex AX2.2.8). Virtually all of the important 

aromatic hydrocarbon precursors emitted in urban atmospheres are lost through reaction with the 

hydroxyl radical. Loss rates for these compounds vary from slow (i.e., benzene) to moderate 

(e.g., toluene), to very rapid (e.g., xylene and trimethylbenzene isomers). These loss rates are 

very well understood at room temperature and atmospheric pressure, and numerous experiments 

have been conducted that verify this. However, the mechanism for the oxidation of aromatic 

hydrocarbons following reaction with OH is poorly understood, as evident from the poor mass 

balance of the reaction products. The mechanism for the oxidation of toluene has been studied 

most thoroughly, and there is general agreement on the initial steps in the mechanism. However, 

at present there is no promising approach for resolving the remaining issues concerning the later 

steps. The oxidation of aromatic hydrocarbons also leads to particle formation which could 

remove gas-phase constituents that participate in O3 formation. What is known of the chemistry 

of secondary organic aerosol formation from gaseous precursors was summarized in the latest 

PM AQCD (U.S. Environmental Protection Agency, 2004). 

The reactions of oxygenated VOCs are also important components of O3 formation 

(Annex AX2.2.9). They may be produced either by the oxidation of hydrocarbons or they may 

be present in ambient air as the result of direct emissions. For example, motor vehicles and 

some industrial processes emit formaldehyde and vegetation emits methanol. 

2-5
As much as 30% of the carbon in hydrocarbons in many urban areas is in the form of 

aromatic compounds. Yet, mass balance analyses performed on irradiated smog chamber 

mixtures of aromatic hydrocarbons indicate that only about one-half of the carbon is in the form 

of compounds that can be identified. The situation is not much better for some smaller 

anthropogenic hydrocarbons. For example, only about 60% of the initial carbon can be 

accounted for in the OH initiated oxidation of 1,3-butadiene. About two-thirds of the initial 

carbon can be identified in product analyses of isoprene oxidation. Adequate analytical 

techniques needed to identify and quantify key intermediate species are not available for many 

compounds. In addition, methods to synthesize many of the suspected intermediate compounds 

are not available so that laboratory studies of their reaction kinetics cannot be performed. 

Similar considerations apply to the oxidation of biogenic hydrocarbons besides isoprene. 

In addition to reactions occurring in the gas phase, reactions occurring on the surfaces of or 

within cloud droplets and airborne particles also occur. Their collective surface area is huge, 

implying that collisions with gas phase species occur on very short time scales. In addition to 

hydrometeors (e.g., cloud and fog droplets and snow and ice crystals) there are also potential 

reactions involving atmospheric particles of varying composition (e.g., wet [deliquesced] 

inorganic particles, mineral dust, carbon chain agglomerates and organic carbon particles) 

to consider. Most of the well-established multiphase reactions tend to reduce the rate of O3 

formation in the polluted troposphere. Removal of HOx and NOx onto hydrated particles will 

reduce the production of O3. However, the photolysis of HONO formed in reactions such as 

these can increase the production of O3. The reactions of Br and Cl containing radicals 

deplete O3 in selected environments such as the Arctic during spring, the tropical marine 

boundary layer and inland salt flats and salt lakes. Direct reactions of O3 and atmospheric 

particles appear to be too slow to reduce O3 formation significantly at typical ambient PM levels. 

In addition, the oxidation of hydrocarbons by Cl radicals could lead to the rapid formation of 

peroxy radicals and higher rates of O3 production in selected coastal environments. It should be 

stressed that knowledge of multiphase processes is still evolving and there are still many 

questions that remain to be answered as outlined in Annex AX2.2.10. 

The oxidants, other than O3, that are formed from the chemistry described above could 

exert effects on human health and perhaps also on vegetation. Gas phase oxidants include 

PAN, H2O2 and CH3OOH and other organic hydroperoxides (Annex AX2.2). In addition to 

2-6
transfer from the gas phase, oxidants can be formed by photochemical reactions occurring in 

particles (Annex 2.2.10.6). However, the pathways leading to the formation of oxidants in the 

particle phase are not as well understood as they are in the gas phase. However, it is to be 

expected that pathways leading to the formation of gas phase oxidants and secondary organic 

aerosols are linked to some degree. In addition, the reaction of O3 with isoprene and other 

biogenic hydrocarbons may also form oxidants in particles. 

Reactions of O3 with monoterpenes have been shown to produce oxidants in the aerosol 

phase. Docherty et al. (2005) found evidence for the substantial production of organic 

hydroperoxides in secondary organic aerosol (SOA) resulting from the reaction of monoterpenes 

with O3. Analysis of the SOA formed in their environmental chamber indicated that the SOA 

was mainly organic hydroperoxides. In particular, they obtained yields of 47% and 85% of 

organic peroxides from the oxidation of "- and $-pinene. The hydroperoxides then react with 

aldehydes in particles to form peroxyhemiacetals, which can either rearrange to form other 

compounds such as alcohols and acids or revert back to the hydroperoxides. The aldehydes are 

also produced in large measure during the ozonolysis of the monoterpenes. Monoterpenes also 

react with OH radicals resulting, however, in the production of more lower molecular weight 

products than in their reaction with O3. Bonn et al. (2004) estimated that hydroperoxides lead to 

63% of global SOA formation from the oxidation of terpenes. The oxidation of anthropogenic 

aromatic hydrocarbons by OH radicals may also produce organic hydroperoxides in SOA 

(Johnson et al., 2004). Although the results of chamber and modeling studies indicate 

substantial production of organic hydroperoxides, it should be noted that data for organic 

hydroperoxides in ambient aerosol samples are sparse. 

Ozone chemical reactions that occur indoors are analogous to those occurring in ambient 

air. In the indoor environment, O3 reacts with unsaturated VOCs, primarily terpenes or terpenerelated compounds from cleaning products, air fresheners, and wood products. The reactions are 

dependent on the O3 indoor concentration, the indoor temperature and, in most cases, the air 

exchange rate/ventilation rate. Some of the reaction products may more negatively impact 

human health and artifacts in the indoor environment than their precursors (Wolkoff et al., 1999; 

Wilkins et al., 2001; Weschler et al., 1992; Weschler and Shields, 1997; Rohr et al., 2002; 

Nøjgaard et al., 2005). Primary reaction products are Criegee biradicals, nitrate radicals, 

and peroxyacetyl radicals. Secondary reaction products are hydroxy, alkyl, alkylperoxy, 

2-7
hydroperoxy, and alkoxy radicals. Reactions with alkenes can produce aldehydes, ketones, 

and organic acids (Weschler and Shields, 2000; Weschler et al., 1992). 

2.3 METEOROLOGICAL PROCESSES AFFECTING OZONE 

Since the 1996 O3 AQCD, substantial new information about transport processes has 

become available from numerical models, field experiments and satellite-based observations. 

Ozone is produced naturally by photochemical reactions in the stratosphere, as shown in Figure 

2-1. Some of this O3 is transported downward into the troposphere throughout the year, with 

maximum contributions during late winter and early spring mainly in a process known as 

tropopause folding. Figure 2-2a shows a synoptic situation associated with a tropopause folding 

event. A vertical cross section taken through the atmosphere from a to aN is shown in Figure 

2-2b. In this figure, the tropopause fold is shown folding downward above and slightly behind 

the surface cold front, bringing stratospheric air with it. Although the tropopause is drawn with 

a solid line, it should not be taken to mean that it is a material surface through which there is no 

exchange. Rather these folds should be thought of as regions in which mixing of tropospheric 

and stratospheric air is occurring (Shapiro, 1980). This imported stratospheric air contributes to 

the natural background of O3 in the troposphere, especially in the free troposphere. It should be 

noted that there is considerable uncertainty in the magnitude and distribution of this potentially 

important source of tropospheric O3. Stratospheric intrusions that reach the surface are rare. 

Much more common are intrusions which penetrate only to the middle and upper troposphere. 

However, O3 transported to the upper and middle troposphere can still affect surface 

concentrations through various exchange mechanisms that mix air from the free troposphere 

with air in the planetary boundary layer. Substantial photochemical production of O3 in the 

troposphere also begins in late winter and early spring; therefore, it cannot be assumed that O3 

present at these times is only stratospheric in origin. The basic atmospheric dynamics and 

thermodynamics of stratospheric-tropospheric exchange are outlined in Annex AX2.3.1. 

Our understanding of the meterological processes associated with summertime O3 episodes 

remains basically the same as outlined in the 1996 O3 AQCD. Major episodes of high O3 

concentrations in the eastern United States and in Europe are associated with slow moving, high 

pressure systems. High pressure systems during the warmer seasons are associated with the 

2-8
Figure 2-2a. Surface weather chart showing sea level (MSL) pressure (kPa), and 

surface fronts. 

Source: Stull (2000). 

Figure 2-2b. Vertical cross section along dashed line (a-aN) from northwest to the 

southeast (CYYC = Calgary, Alberta; LBF = North Platte, NB; LCH = Lake 

Charles, LA). The approximate location of the jet stream core is indicated 

by the hatched area. The position of the surface front is indicated by the 

cold-frontal symbols and the frontal inversion top by the dashed line. 

Note: This is 12 h later than the situations shown in Figure 2-2a. 

Source: Adapted from Stull (2000). 

2-9 
sinking of air, resulting in warm, generally cloudless skies, with light winds. The sinking of air 

results in the development of stable conditions near the surface which inhibit or reduce the 

vertical mixing of O3 precursors. The combination of inhibited vertical mixing and light winds 

minimizes the dispersal of pollutants emitted in urban areas, allowing their concentrations to 

build up. Photochemical activity involving these precursors is enhanced because of higher 

temperatures and the availability of sunlight. In the eastern United States, high O3 

concentrations during a large scale episode can extend over hundreds of thousands of square 

kilometers for several days. These conditions have been described in greater detail in the 

1996 O3 AQCD. The transport of pollutants downwind of major urban centers is characterized 

by the development of urban plumes. However, the presence of mountain barriers limits mixing 

(as in Los Angeles and Mexico City) and results in a higher frequency and duration of days with 

high O3 concentrations. Ozone concentrations in southern urban areas (such as Houston, TX and 

Atlanta, GA) tend to decrease with increasing wind speed. In northern cities (such as Chicago, 

IL; New York, NY; Boston, MA; and Portland, ME), the average O3 concentrations over the 

metropolitan areas increase with wind speed, indicating that transport of O3 and its precursors 

from upwind areas is important (Husar and Renard, 1998; Schichtel and Husar, 2001). 

Ozone and other secondary pollutants are determined by meteorological and chemical 

processes extending typically over spatial scales of several hundred kilometers (e.g., Civerolo 

et al., 2003; Rao et al., 2003). An analysis of the output of regional model studies conducted by 

Kasibhatla and Chameides (2000) suggests that O3 can be transported over a few thousand 

kilometers in the upper boundary layer of the eastern half of the United States during some O3 

episodes. Convection is capable of transporting O3 and its precursors vertically through the 

troposphere, as shown in Annex AX2.3.2. Nocturnal low level jets (LLJs) can also transport 

pollutants hundreds of kilometers (Annex AX2.3.3). Schematic diagrams showing the 

atmospheric conditions during the formation of LLJs and the regions in which they are most 

prevalent are given in Figures 2-3 and 2-4. Such LLJs have also been observed off the coast of 

California. Turbulence associated with LLJs can bring these pollutants to the surface and result 

in secondary O3 maxima during the early morning in many locations (Corsmeier et al., 1997). 

Aircraft observations indicate that there can be substantial differences in mixing ratios of 

key species between the surface and the atmosphere above (Fehsenfeld et al., 1996; Berkowitz 

and Shaw, 1997). In particular, mixing ratios of O3 can be higher in the lower free troposphere 

2-10
Figure 2-3. The diurnal evolution of the planetary boundary layer (PBL) while high 

pressure prevails over land. Three major layers exist (not including the 

surface layer): a turbulent mixed layer; a less turbulent residual layer which 

contains former mixed layer air; and a nocturnal, stable boundary layer 

that is characterized by periods of sporadic turbulence. 

Source: Adapted from Figures 1.7 and 1.12 of Stull (1999). 

(aloft) than in the planetary boundary layer (PBL) during multiday O3 episodes (Taubmann et al., 

2004, 2005). These conditions are illustrated schematically in Figure 2-5. Convective processes 

and small scale turbulence transport O3 and other pollutants both upward and downward 

throughout the planetary boundary layer and the free troposphere. Ozone and its precursors can 

be transported vertically by convection into the upper part of the mixed layer on one day, then 

transported overnight as a layer of elevated mixing ratios, and then entrained into a growing 

convective boundary layer downwind and brought back down to the surface. High O3 

concentrations showing large diurnal variations at the surface in southern New England were 

associated with the presence of such layers (Berkowitz et al., 1998). Because of wind shear, 

winds several hundred meters above the ground can bring pollutants from the west, even though 

surface winds are from the southwest during periods of high O3 in the eastern United States 

(Blumenthal et al., 1997). These considerations suggest that, in many areas of the United States, 

O3 formation involves processes occurring over hundreds if not thousands of square kilometers. 

2-11
Figure 2-4. Locations of low level jet occurrences in decreasing order of prevalence 

(most frequent, common, observed). These locations are based on 2-years 

radiosonde data obtained over limited areas. With better data coverage, 

other low level jets may well be observed elsewhere in the United States. 

Source: Bonner (1968). 

Although the vast majority of measurements are made near the Earth’s surface, there is 

substantial photochemistry and transport of O3 occurring above the boundary layer in the free 

troposphere. In the free troposphere, pollutants are chemically more stable and can be 

transported over much longer distances and O3 is produced more efficiently than in the planetary 

boundary layer. Results from the Atmosphere/Ocean Chemistry Experiment (AEROCE) 

indicated that springtime maxima for surface O3 over the western North Atlantic Ocean result 

from tropopause folding in close proximity to convective clouds (Annex AX2.3.4). The 

convection lifts O3 and its precursors to the free troposphere where they mix with O3 from the 

stratosphere and the mixture is transported eastward. Results from the North Atlantic Regional 

Experiment (Annex AX2.3.4) indicate that summertime air is transported along the East Coast 

northeastward and upward ahead of cold fronts. New England and the Maritime Provinces of 

Canada receive substantial amounts of O3 and other pollutants through this mechanism. 

2-12
PM

Figure 2-5. Conceptual two-reservoir model showing conditions in the planetary 

boundary layer (PBL) and in the lower free troposphere during a multiday O3 

episode. The dotted line represents the top of PBL. Emissions occur in the 

PBL, where small, unmixed black carbon, sulfate, and crustal particles in the 

2.5 size range are also shown. Ozone concentrations as well as potential 

temperature (2) and actinic flux are lower in the PBL than in the lower free 

troposphere, while relative humidity and the Angstrom exponent for 

aerosol scattering (") are higher. Larger, internally mixed sulfate and 

carbonaceous particles (still in the PM2.5 size range) and more O3 exist in the 

lower free troposphere. 

Source: Taubman et al. (2004, 2005). 

Pollutants transported in this way can then be entrained in stronger and more stable westerly 

winds aloft and can travel across the North Atlantic Ocean. The pollutants can then be brought 

to the surface by subsidence in high pressure systems (typically behind the cold front in advance 

of the one mentioned above). Thus, pollutants from North America can be brought down either 

over the North Atlantic Ocean or in Europe. Pollutants can be transported across the North 

Pacific Ocean from Asia to North America in a similar way. Behind an advancing cold front, 

cold and dry stratospheric air is also being transported downward and southward. Stratospheric 

2-13
constituents and tropospheric constituents can then mix by small-scale turbulent exchange 

processes. The results of these studies suggest that the mechanisms involved in the long-range 

transport of O3 and its precursors are closely tied to the processes involved in stratospherictropospheric exchange. Land-sea breezes affect the concentration and dispersal of pollutants in 

coastal zone cities (Annex AX2.5). 

The local rate of O3 formation depends on atmospheric conditions such as the availability 

of solar ultraviolet radiation capable of initiating photolysis reactions, air temperatures and the 

concentrations of chemical precursors (Annex AX2.3.6). The dependence of daily maximum 

8-h O3 concentrations on daily maximum temperature is illustrated in Figure 2-6 for the 

Baltimore, MD area. As can be seen, O3 concentrations tend to increase with temperature 

(r = 0.74). However, this trend is absent in data from Phoenix, AZ as can be seen in Figure 2-7 

(r = 0.14). These figures show that relations of O3 to precursor variables are location-specific 

and relations observed in one area cannot be readily extrapolated to another. Factors that may be 

responsible for the differences in O3 behavior in the two areas are discussed in Section AX2.3.6. 

Figure 2-6. A scatter plot of daily maximum 8-h average O3 concentrations versus daily 

maximum temperature for May through September 1994 to 2004 in the 

Baltimore, MD Air Quality Forecast Area. 

Source: Piety (2005). 

2-14 
Figure 2-7. A scatter plot of daily maximum 8-h average O3 concentrations versus daily 

maximum temperature for May through September 1996 to 2004 at sites 

downwind of Phoenix, AZ. 

Source: Piety (2005). 

2.4 RELATIONS OF OZONE TO ITS PRECURSORS 

Rather than varying directly with emissions of its precursors, O3 changes in a nonlinear 

fashion with the concentrations of its precursors (Annex AX2.4). At the low NOx concentrations 

found in most environments, ranging from remote continental areas to rural and suburban areas 

downwind of urban centers (low - NOx regime), the net production of O3 increases with 

increasing NOx. At the high NOx concentrations found in downtown metropolitan areas, 

especially near busy streets and roadways and in power plant plumes, there is scavenging 

(titration) of O3 by reaction with NO (high - NOx regime). In between these two regimes, there is 

a transition stage in which O3 shows only a weak dependence on NOx concentrations. In the 

high - NOx regime, NO2 scavenges OH radicals which would otherwise oxidize VOCs to 

produce peroxy radicals, which in turn would oxidize NO to NO2. In this regime, O3 production 

is limited by the availability of free radicals. The production of free radicals is in turn limited by 

the availability of solar UV radiation capable of photolyzing O3 (in the Hartley bands) or 

2-15
aldehydes and/or by the abundance of VOCs whose oxidation produce more radicals than they 

consume. In the low-NOx regime, the overall effect of the oxidation of VOCs is to generate 

(or at least not consume) free radicals, and O3 production varies directly with NOx. There are a 

number of ways to refer to the chemistry in these two chemical regimes. Sometimes the terms 

VOC-limited and NOx-limited are used. However, there are difficulties with this usage because 

(1) VOC measurements are not as abundant as they are for nitrogen oxides, (2) rate coefficients 

for reaction of individual VOCs with free radicals vary over an extremely wide range, and 

(3) consideration is not given to CO nor to reactions that can produce free radicals without

involving VOCs. The terms NOx-limited and NOx-saturated (e.g., Jaeglé et al., 2001) will be 

used wherever possible to more adequately describe these two regimes. However, the 

terminology used in original articles will also be used here. 

The chemistry of OH radicals, which are responsible for initiating the oxidation of 

hydrocarbons, shows behavior similar to that for O3 with respect to NOx concentrations 

(Hameed et al., 1979; Pinto et al., 1993; Poppe et al., 1993; Zimmerman and Poppe, 1993). 

These considerations introduce a high degree of uncertainty into attempts to relate changes in O3 

concentrations to emissions of precursors. There are no definitive rules governing the levels 

of NOx at which the transition from NOx-limited to NOx-saturated conditions occurs. The 

transition between these two regimes is highly spatially and temporally dependent and depends 

also on the nature and abundance of the hydrocarbons that are present. 

Trainer et al. (1993) and Olszyna et al. (1994) have shown that O3 and NOy are highly 

correlated in rural areas in the eastern United States. Trainer et al. (1993) also showed that O3 

levels correlate even better with NOz than with NOy, as may be expected because NOz represents 

the amount of NOx that has been oxidized, forming O3 in the process. NOz is equal to the 

difference between measured total reactive nitrogen (NOy) and NOx and represents the summed 

products of the oxidation of NOx. NOz is composed mainly of HNO3, PAN and other organic 

nitrates, particulate nitrate, and HNO4. 

Trainer et al. (1993) also suggested that the slope of the regression line between O3 

and NOz can be used to estimate the rate of O3 production per NOx oxidized (also known as 

the O3 production efficiency, or OPE). Ryerson et al. (1998, 2001) used measured correlations 

between O3 and NOz to identify different rates of O3 production in plumes from large point 

sources. A number of studies in the planetary boundary layer over the continental United States 

2-16
have found that the OPE ranges typically from one to nearly ten. However, it may be higher in 

the upper troposhere and in certain areas, such as the Houston-Galveston area. Observations 

indicate that the OPE depends mainly on the abundance of NOx. 

Various techniques have been proposed to use ambient NOx and VOC measurements to 

derive information about the dependence of O3 production on their concentrations. For example, 

it has been suggested that O3 formation in individual urban areas could be understood in terms of 

measurements of ambient NOx and VOC concentrations during the early morning (e.g., National 

Research Council, 1991). In this approach, the ratio of summed (unweighted) VOC to NOx is 

used to determine whether conditions were NOx-limited or VOC limited. This procedure is 

inadequate because it omits many factors that are important for O3 production such as the impact 

of biogenic VOCs (which are typically not present in urban centers during early morning); 

important differences in the ability of individual VOCs to generate free radicals (rather than just 

total VOC) and other differences in O3 forming potential for individual VOCs (Carter, 1995); 

and changes in the VOC to NOx ratio due to photochemical reactions and deposition as air 

moves downwind from urban areas (Milford et al., 1994). 

Photochemical production of O3 generally occurs simultaneously with the production of 

various other species such as nitric acid (HNO3), organic nitrates, and other oxidants such as 

hydrogen peroxide. The relative rate of production of O3 and other species varies depending on 

photochemical conditions, and can be used to provide information about O3-precursor 

sensitivity. Sillman (1995) and Sillman and He (2002) identified several secondary reaction 

products that show different correlation patterns for NOx-limited and NOx-saturated conditions. 

The most important correlations are for O3 versus NOy, O3 versus NOz, O3 versus HNO3, 

and H2O2 versus HNO3. The correlations between O3 and NOy, and O3 and NOz are especially 

important because measurements of NOy and NOx are more widely available than for VOCs. 

Measured O3 versus NOz (Figure 2-8) shows distinctly different patterns in different locations. 

In rural areas and in urban areas such as Nashville, TN, O3 is highly correlated with NOz. 

By contrast, in Los Angeles, CA, O3 is not as highly correlated with NOz, and the rate of increase 

of O3 with NOz is lower and the O3 concentrations for a given NOz value are generally lower. 

The different O3 versus NOz relations in Nashville, TN and Los Angeles, CA reflects the 

difference between NOx-limited conditions in Nashville versus an approach to NOx- saturated 

conditions in Los Angeles. 

2-17
Figure 2-8. Measured values of O3 and NOz (NOy – NOx) during the afternoon at rural 

sites in the eastern United States (grey circles) and in urban areas and urban 

plumes associated with Nashville, TN (gray dashes); Paris, France (black 

diamonds); and Los Angeles CA (Xs). 

Sources: Trainer et al. (1993), Sillman et al. (1997, 1998), Sillman and He (2002). 

The difference between NOx-limited and NOx-saturated regimes is also reflected in 

measurements of hydrogen peroxide (H2O2). Hydrogen peroxide production is highly sensitive 

to the abundance of free radicals and is thus favored in the NOx-limited regime. Measurements 

in the rural eastern United States (Jacob et al., 1995), Nashville, TN (Sillman et al., 1998), and 

Los Angeles, CA (Sakugawa and Kaplan, 1989), show large differences in H2O2 concentrations 

between likely NOx-limited and NOx-saturated locations. 

2.5 THE ROLE OF CHEMISTRY-TRANSPORT MODELS IN 

UNDERSTANDING ATMOSPHERIC OZONE 

Chemistry-transport models (CTMs) are used to improve understanding of atmospheric 

chemical processes and to develop control strategies (Annex AX2.5). The main components of a 

CTM are summarized in Figure 2-9. Models such as the CMAQ (Community Model for Air 

Quality) system incorporate numerical algorithms describing the processes shown in Figure 2-9. 

Also shown in Figure 2-9 is the meteorological model used to provide the inputs for calculating 

2-18
Figure 2-9. Main components of a comprehensive atmospheric chemistry modeling 

system, such as Models-3. 

the transport of species in the CTM. Meteorological models, such as MM5, which supply these 

inputs to the CTMs mentioned above, also provide daily weather forecasts. The domains of 

these models extend typically over areas of millions of square kilometers. 

Because these models are computationally intensive, it is often impractical to run them 

over larger domains without sacrificing some features. For these reasons, both the 

meteorological model and the CTM rely on boundary conditions that allow processes occurring 

outside the model domain to influence their predictions. The entire system, consisting of 

meteorological model, emissions processor, and output processors shown in Figure 2-9 

constitutes the framework of EPA’s Models-3. 

2-19
Because of the large number of chemical species and reactions that are involved in the 

oxidation of realistic mixtures of anthropogenic and biogenic hydrocarbons, condensed 

mechanisms must be used in atmospheric models. These mechanisms are tested by comparison 

with smog chamber data. However, the existing chemical mechanisms often neglect many 

important processes such as the formation and subsequent reactions of long-lived carbonyl 

compounds, the incorporation of the most recent information about intermediate compounds, and 

heterogeneous reactions involving cloud droplets and aerosol particles. 

Emissions inventories are compiled for O3 precursors ( NOx, VOCs, and CO). Recent 

estimates and more detailed discussions of the estimates are given in Annex AX2.5.2. 

Anthropogenic NOx emissions are associated with combustion processes. Most emissions are in 

the form of NO, which is formed at high combustion temperatures from atmospheric nitrogen 

and oxygen and from fuel nitrogen. The two largest sources of NOx are electric power 

generation plants and motor vehicles. Emissions of NOx therefore are highest in areas having a 

high density of power plants and in urban regions having high traffic density. Natural NOx 

sources include stratospheric intrusions, lightning, soils, and wildfires. Lightning, fertilized 

soils, and wildfires are the major natural sources of NOx in the United States. Both nitrifying 

and denitrifying organisms in the soil can produce NOx, mainly in the form of NO. Emission 

rates depend mainly on fertilization levels and soil temperature and moisture. Spatial and 

temporal variability in soil NOx emissions leads to considerable uncertainty in emissions 

estimates. Nationwide, about 60% of lightning generated NOx occurs in the southern 

United States and about 60% the total NOx emitted by soils occurs in the central corn belt of 

the United States. The oxidation of NH3 emitted mainly by livestock and soils, leads to the 

formation of a small amount of NO. Uncertainties in natural NOx inventories are much larger 

than for anthropogenic NOx emissions. 

Hundreds of VOCs, containing mainly two to about twelve carbon atoms, are emitted by 

evaporation and combustion processes from a large number of anthropogenic sources. The two 

largest anthropogenic source categories in the U.S. EPA’s emissions inventories are industrial 

processes and transportation. Emissions of VOCs from highway vehicles account for roughly 

two-thirds of the transportation-related emissions. 

The accuracy of VOC emission estimates is difficult to determine, both for stationary and 

mobile sources. Evaporative emissions, which depend on temperature and other environmental 

2-20
factors, compound the difficulties of assigning accurate emission factors. In assigning VOC 

emission estimates to the mobile source category, models are used that incorporate numerous 

input parameters (e.g., type of fuel used, type of emission controls, age of vehicle), each of 

which has some degree of uncertainty. Data for the ratio of CO to NOx and NMHC to NOx in 

traffic tunnels (e.g., Pierson et al., 1990) indicated that emissions of NMHCs and CO from motor 

vehicles have been underestimated by as much as a factor of two (based on the assumption that 

emissions of NOx were reasonably well represented in the inventories). However, the results of 

more recent studies have been mixed, with many studies showing agreement to within ±50%, as 

summarized in Air Quality Criteria for Carbon Monoxide (U.S. Environmental Protection 

Agency, 2000). Remote sensing data (Stedman et al., 1991) indicate that about 50% of NMHC 

and CO emissions are produced by about 10% of the vehicles. These “super-emitters” are 

typically poorly maintained. Vehicles of any age engaged in off-cycle operations (e.g., rapid 

accelerations) emit much more than if operated in normal driving modes. 

Vegetation emits significant quantities of VOCs, such as terpenoid compounds (isoprene, 

2-methyl-3-buten-2-ol, monoterpenes), compounds in the hexanal family, alkenes, aldehydes, 

organic acids, alcohols, ketones, and alkanes. The major chemicals emitted by plants are 

isoprene (35%), 19 other terpenoid compounds and 17 non-terpenoid compounds including 

oxygenated compounds (40%) (Guenther et al., 2000). Coniferous forests represent the largest 

source on a nationwide basis, because of their extensive land coverage. Most biogenic emissions 

occur during the summer, because of their dependence on temperature and incident sunlight. 

Biogenic emissions are also higher in southern states than in northern states for these reasons and 

because of species variations. The uncertainty in natural emissions is about 50% for isoprene 

under midday summer conditions and could be as much as a factor of ten higher for some 

compounds (Guenther et al., 2000). Uncertainties in both biogenic and anthropogenic VOC 

emission inventories prevent determination of the relative contributions of these two categories 

at least in many urban areas. On the regional and global scales, emissions of VOCs from 

vegetation are much larger than those from anthropogenic sources. 

The performance of CTMs must be evaluated by comparison with field data as part of a 

cycle of model evaluations and subsequent improvements. Evaluations of the CMAQ are given 

in Arnold et al. (2003) and Fuentes and Raftery (2005). Discrepancies between model 

predictions and observations can be used to point out gaps in current understanding of 

2-21
atmospheric chemistry and to spur improvements in parameterizations of atmospheric chemical 

and physical processes. Model evaluation does not merely involve a straightforward comparison 

between model predictions and the concentration field of the pollutant of interest. Such 

comparisons may not be meaningful because it is difficult to determine if agreement between 

model predictions and observations truly represents an accurate treatment of physical and 

chemical processes in the CTM or the effects of compensating errors in complex model routines. 

Ideally, each of the model components (emissions inventories, chemical mechanism, 

meteorological driver) should be evaluated individually. However, this is rarely done in 

practice. A comparison between free radical concentrations predicted by parameterized 

chemical mechanisms and observations suggests that radical concentrations were overestimated 

by current chemical mechanisms for NOx concentrations <~5 ppb (Volz-Thomas et al., 2003). 

In addition to comparisons between concentrations of calculated and measured species, 

comparisons of correlations between measured primary VOCs and NOx and modeled VOCs 

and NOx are especially useful for evaluating results from chemistry-transport models. Likewise, 

comparisons of correlations between measured species and modeled species can be used to 

provide information about the chemical state of the atmosphere and to evaluate model 

representations (including O3 production per NOx, O3-NOx-VOC sensitivity, and the general 

accuracy of photochemical representations). A CTM that demonstrates the accuracy of both its 

computed VOC and NOx in comparison with ambient measurements and the spatial and temporal 

relations among the critical secondary species associated with O3 has a higher probability of 

representing O3-precursor relations correctly than one that does not. 

2.6 TECHNIQUES FOR MEASURING OZONE AND ITS PRECURSORS 

Several techniques have been developed for sampling and measurement of O3 in the 

ambient atmosphere at ground level. Although the chemiluminescence method (CLM) using 

ethylene is designated as the Federal Reference Method (FRM) for measuring O3, monitoring in 

the NAMS/SLAMS networks is conducted mainly with UV absorption spectrometry using 

commercial short path instruments. The primary reference standard instrument is a relatively 

long-path UV absorption spectrometer maintained under carefully controlled conditions at NIST 

(e.g., Fried and Hodgeson, 1982). Episodic measurements are made with a variety of other 

2-22
techniques based on the principles of chemiluminescence, electrochemistry, differential optical 

absorption spectroscopy (DOAS), and LIDAR. 

In principle, each of these methods is subject to interference. Kleindienst et al. (1993) 

found that water vapor could cause a positive interference in the CLM with an average positive 

deviation of 3% ozone/% water vapor at 25 °C. However, they also noted that water vapor could 

cause positive interferences of up to 9% at high humidities (dew point of 24 °C). The UV 

absorption spectrometers are subject to positive interference by atmospheric constituents, such as 

certain aromatic aldehydes that absorb at the 253.7 nm Hg resonance line and are at least 

partially removed by the MnO2 scrubber. Parrish and Fehsenfeld (2000) did not find any 

evidence for significant interference (>1%) in flights through the Nashville urban plume. The 

same group tested the air of Houston, El Paso, Nashville, Los Angeles, San Francisco and the 

East Coast. They observed only one instance of substantive positive interference defined as the 

UV absorption technique showing more than a few ppb more than the CLM. This occurred in 

Laporte, TX under heavily polluted conditions and a low inversion, at night (Jobson et al., 2004). 

Leston et al. (2005) observed interference of from 20 to 40 ppb in Mexico City and in a separate 

smog chamber study. However, the concentrations of relevant compounds were many times 

higher than found in U.S. urban areas. Thus, it is not likely that such interference could be more 

than a few ppb under typical ambient conditions. However, Leston et al. (2005) suggested that 

the use of other materials in the scrubber could have eliminated the interference seen in their 

smog chamber study. 

By far, most measurements of NO are made using the CLM, based on its reaction with O3. 

Commercial instruments for measuring NO and NO2 are constructed with an internal converter 

for reducing NO2 to NO and then measuring NO by the CLM. In principle, this technique yields 

a measurement of NOx with NO2 found by difference between NOx and NO; but, these 

converters also reduce NOz compounds thereby introducing a positive interference in the 

measurement of NO2. Other methods for measuring NO2 are available, such as photolytic 

reduction followed by CLM, laser-induced fluorescence and DOAS. However, they require 

further development before they can be used for routine monitoring in the NAMS/SLAMS 

networks. More detailed descriptions of the issues and techniques discussed above and 

techniques for measuring HNO3 and VOCs can be found in Annex AX2.6. 

2-23
2.7 SUMMARY 

Ozone is formed by atmospheric reactions involving two classes of precursor compounds, 

volatile organic compounds (VOCs) and nitrogen oxides (NOx). Ozone is thus a secondary 

pollutant. Ozone is ubiquitous throughout the atmosphere; it is present even in remote areas of 

the globe. The photochemical oxidation of almost all anthropogenic and biogenic VOCs is 

initiated by reaction with hydroxyl (OH) radicals. At night, when they are most abundant, NO3 

radicals also oxidize alkenes. In coastal and other select environments, Cl and Br radicals can 

also initiate the oxidation of VOCs. 

In urban areas, basically all classes of VOCs (alkanes, alkenes, aromatic hydrocarbons, 

carbonyl compounds, etc.) and CO are important for O3 formation. Although knowledge of the 

oxidative mechanisms of VOCs has improved over the past several years, gaps in knowledge 

involving key classes, such as aromatic hydrocarbons, still remain. For example, only about half 

of the carbon initially present in aromatic hydrocarbons in smog chamber studies form 

compounds that have been identified. 

In addition to gas phase reactions, reactions also occur on the surfaces of, or within cloud 

droplets and airborne particles. Most of the well-established multiphase reactions tend to reduce 

the rate of O3 formation in polluted environments. Reactions of Cl and Br containing radicals 

deplete O3 in selected environments such as the Arctic during spring, the tropical marine 

boundary layer and inland salt lakes. Direct reactions of O3 with atmospheric particles appear to 

be too slow to reduce O3 formation significantly at typical ambient PM levels. 

Our basic understanding of the meteorological processes associated with summertime O3 

episodes has not changed over the past several years. However, the realization that long-range 

transport processes are important for determining O3 concentrations at the surface is growing. 

In addition to synoptic scale flow fields, nocturnal low-level jets are capable of transporting 

pollutants hundreds of km from their sources in either the upper boundary layer or the lower free 

troposphere. Turbulence then brings O3 and other pollutants to the surface. On larger scales, 

important progress has been made in identifying the mechanisms of intercontinental transport 

of O3 and other pollutants. 

Some O3 would be found near the earth’s surface as the result of its downward transport 

from the stratosphere, even in the absence of photochemical reactions in the troposphere. 

Intrusions of stratospheric O3 that reach the surface are rare. Much more common are intrusions 

2-24
that penetrate to the middle and upper troposphere. However, O3 transported to the middle and 

upper troposphere can still affect surface concentrations through various mechanisms that mix 

air between the planetary boundary layer and the free troposphere above. 

The formation of O3 and associated compounds is a complex, nonlinear function of many 

factors, including the intensity and spectral distribution of sunlight; atmospheric mixing and 

other atmospheric processes; and the concentrations of the precursors in ambient air. At lower 

NOx concentrations found in most environments, ranging from remote continental areas to rural 

and suburban areas downwind of urban centers, the net production of O3 increases with 

increasing NOx. At higher concentrations found in downtown metropolitan areas, especially 

near busy streets and highways and in power plant plumes, there is net destruction of O3 by 

reaction with NO. In between these two regimes, there is a transition stage in which O3 

production shows only a weak dependence on NOx concentrations. The efficiency of O3 

production per NOx oxidized is generally highest in areas where NOx concentrations are lowest 

and decrease with increasing NOx concentration. 

Chemistry transport models are used to improve understanding of atmospheric chemical 

and physical processes as well as to develop air pollution control strategies. The performance of 

these models must be evaluated by comparison with field data as part of a cycle of model 

evaluations and subsequent improvements. Discrepancies between model predictions and 

observations can be used to point out gaps in current understanding and thus to improve 

parameterizations of atmospheric chemical and physical processes. Model evaluation does not 

merely involve a straightforward comparison between model predictions and the concentration 

fields of a pollutant of interest (e.g., O3). Such comparisons may not be meaningful because it is 

difficult to determine if agreement between measurements and model predictions truly represents 

an accurate treatment of physical and chemical processes in the model or the effects of 

compensating errors in model routines. 

The main methods currently in use for routine monitoring of ambient O3 are based on 

chemiluminescence or UV absorption. Measurements at most ambient monitoring sites are 

based on UV absorption. Both of these methods are subject to interference by other atmospheric 

components. One study found large positive interference in Mexico City and in a smog 

chamber, but a few studies conducted in urban plumes did not find significant positive 

interference in the UV absorption technique. Sufficient new information is not available to 

2-25
amend estimates of the accuracy and precision of O3 monitors. Such a reevaluation requires 

studies of the simultaneous effects of a number of potential interferants including water vapor, 

organic compounds, and temperature on the UV photometric and chemiluminescent methods. 

2-26
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3. ENVIRONMENTAL CONCENTRATIONS,

PATTERNS, AND EXPOSURE ESTIMATES

3.1 INTRODUCTION 

Identification and Use of Existing Air Quality Data 

Topics discussed in this chapter include the characterization of ambient air quality data for 

ozone (O3), the uses of these data in assessing the exposure of vegetation to O3, concentrations 

of O3 in microenvironments, and a discussion of the currently available human exposure data and 

exposure model development. The information contained in this chapter pertaining to ambient 

concentrations is taken primarily from the U.S. Environmental Protection Agency (EPA) 

Air Quality System (AQS; formerly the AIRS database). The AQS contains readily accessible 

detailed, hourly data that has been subject to EPA quality control and assurance procedures. 

Data available in AQS were collected from 1979 to 2001. As discussed in the 1996 O3 Air 

Quality Criteria Document or AQCD (U.S. Environmental Protection Agency, 1996), the data 

available prior to 1979 may be unreliable due to calibration problems and other uncertainties. 

As noted in the 1996 O3 AQCD (U.S. Environmental Protection Agency, 1996), O3 is the 

only photochemical oxidant other than nitrogen dioxide (NO2) that is routinely monitored and 

for which a comprehensive database exists. Data for peroxyacetyl nitrate (PAN), hydrogen 

peroxide (H2O2), and other oxidants either in the gas phase or particle phase typically have been 

obtained only as part of special field studies. Consequently, no data on nationwide patterns of 

occurrence are available for these non-O3 oxidants; nor are extensive data available on the 

relationships of levels and patterns of these oxidants to those of O3. However, available data for 

gas phase and particle phase oxidants are discussed in this chapter. 

Characterizing Ambient Ozone Concentrations 

The “concentration” of a specific air pollutant is typically defined as the amount (mass) 

of that material per unit volume of air. However, most of the data presented in this chapter are 

expressed as “mixing ratios” in terms of a volume-to-volume ratio (parts per million [ppm] or 

parts per billion [ppb]). Data expressed this way are often referred to as concentrations, both in 

3-1
the literature and in the text, following common usage. Human exposures are expressed in units 

of mixing ratio times time. 

Several different air quality metrics have been suggested for evaluating exposures of 

vegetation to O3. The peak-weighted, cumulative exposure indicators used in this chapter for 

characterizing vegetation exposures are SUM06 and SUM08 (the sums of all hourly average 

concentrations $0.06 and 0.08 ppm, respectively) and W126 (the sum of the hourly average 

concentrations that have been weighted according to a sigmoid function that is based on a 

hypothetical vegetation response [see Lefohn and Runeckles, 1987]). Further discussion of these 

exposure indices is presented in Chapter 9. 

The EPA has established “ozone seasons” during which measurement of ambient O3 

concentrations for different locations within the United States and the U.S. territories is required 

(CFR, 2000). Table AX3-1 shows the O3 seasons during which continuous, hourly averaged O3 

concentrations must be monitored. Monitoring is optional outside of these O3 seasons and 

indeed is conducted during the winter in a number of areas. 

Data for O3 in ambient air across the United States are summarized in Section 3.2. The 

data are summarized for urban, rural, and relatively remote sites. Relatively remote monitoring 

sites (RRMS) are sites that are not strongly influenced by nearby pollution sources and are 

located mainly in national parks in the West. However, this does not mean that they are free of 

the effects of regional or local pollution, especially during tourist seasons. Data for the spatial 

variability of O3 within urban areas are summarized in Section 3.3. Data for the diurnal and 

seasonal variability of O3 concentrations are given in Section 3.4. The long term temporal 

variability of O3 concentrations is discussed in Section 3.5. Relationships among O3 and other 

species are discussed in Section 3.6. Information about the occurrence of other oxidants and 

their relationship to O3 is given in this section. A discussion of Policy Relevant Background 

(PRB) O3 concentrations is presented in Section 3.7. PRB O3 concentrations are background O3 

concentrations that would be observed in the U.S. in the absence of anthropogenic emissions 

of O3 precursors in the U.S., Canada, and Mexico. Background levels so defined help facilitate 

the distinction between pollution levels that can be controlled by U.S. regulations (or through 

international agreements with neighboring countries) from levels that are generally 

uncontrollable by the U.S. Indoor sources and emissions of O3 are discussed in Section 3.8. 

3-2
Issues related to evaluating human exposure to O3 are summarized in Section 3.9. Finally, 

a summary of key points in Chapter 3 is given in Section 3.10. 

3.2 AMBIENT AIR QUALITY DATA FOR OZONE 

Ozone Air Quality at Urban, Suburban, and Nonurban Sites 

Figure 3-1 shows the mean daily maximum 8-h O3 concentrations, and Figure 3-2 shows 

the 95th percentile values of the daily maximum 8-h O3 concentrations, based on countywide 

site-wise averages across the United States from May to September 2000 to 2004. The period 

from May to September was chosen because, although O3 is monitored for different lengths of 

time across the country, all O3 monitors should be operational during these months. Data 

flagged because of quality control issues were removed with concurrence of the local monitoring 

agency. Only days with data for 18 of 24 hours were kept, and a minimum of 115 of 153 days 

were required in each year. Cut points for the tertile distributions on each map were chosen at 

the median and 95th percentile values. These cut points were chosen because they represent 

standard metrics for characterizing important aspects of human exposure used by the EPA. Any 

other percentiles or statistics that are believed to be helpful for characterizing human exposures 

could also be used. Blank areas on the maps indicate no data coverage. It should be noted that 

county areas can be much larger in the West than in the East, but monitors are not spread evenly 

within a county. As a result, the assigned concentration range might not represent conditions 

throughout a particular county and, so, large areas in western counties where there are no 

monitors were blanked out. 

As shown in Figure 3-1, the median of the countywide, mean daily maximum 8-h O3 

concentration across the United States is 49 ppb, and 5% of these site-wise means exceeded 

57 ppb. Though the median and 95th percentile values are fairly close, these results cannot be 

taken to imply that average O3 concentrations lie in a relatively narrow range throughout the 

United States, because data coverage is not as complete in the West as it is in the East. High 

mean daily maximum 8-h O3 concentrations are found in California and states in the Southwest 

as well as in several counties in the East. As shown in Figure 3-2, the nationwide median of the 

countywide, 95th percentile value of the daily maximum 8-h O3 concentration is 73 ppb and 5% 

of these values are above 85 ppb. High values for the 95th percentiles are found in California, 

3-3
Figure 3-1. Countywide mean daily maximum 8-h O3 concentrations, May to September 

2000 to 2004. 

Source: Fitz-Simons et al. (2005). 

Texas, and some counties in the East, but not necessarily in the same counties in the East as 

shown for the mean daily maximum 8-h concentrations in Figure 3-1. 

Although mean O3 concentrations in Houston, TX were below the nationwide median, its 

95th percentile value ranks in the highest 5% nationwide. Conversely, mean O3 concentrations 

in southwestern states are among the highest in the United States, but values at the upper end of 

the distribution (e.g., the 95th percentile value) in these states are not among the highest peak 

values in the United States. In other areas where the highest mean O3 concentrations occurred 

(e.g., California; Dallas-Fort Worth, TX; and the Northeast Corridor), the highest peak values 

were also observed. 

3-4
Figure 3-2. Countywide 95th percentile value of daily maximum 8-h O3 concentrations, 

May to September 2000 to 2004. 

Source: Fitz-Simons et al. (2005). 

Although countywide averages are shown, it should be noted that considerable spatial 

variability can exist within a county, especially within urban areas as described in Section 3.3. 

In addition, there can also be differences in the diurnal profile of O3 among monitors within 

counties. 

Box plots showing the distribution of nationwide O3 concentrations for different averaging 

periods (1-h daily maximum, 8-h daily maximum, and 24-h daily average) are given in Figures 

AX3-4 to AX3-6 and numerical values are given in Table AX3-2. The differences between the 

50th and 95th percentile values indicate the range of O3 levels between “typical” O3 days and 

“high” O3 days. These differences are approximately 40, 30, and 25 ppb for the daily 1-h and 

3-5
8-h daily maxima and 24-h average O3 concentrations, respectively. As might be expected, the 

daily maximum 1-h and 8-h O3 concentrations are highly correlated. 

Lehman et al. (2004) divided the eastern United States into five regions, each of which 

exhibit relatively distinct spatial and temporal patterns of O3 concentrations at nonurban sites. 

Only sites classified as being rural or suburban and with land usage of forest, agriculture, or 

residential were included in the analyses. These criteria were chosen to avoid sites where O3 is 

scavenged by NO, which can be found in high concentrations near major sources such as traffic 

in urban cores. The five regions, shown in Figure 3-3, are characterized by different patterns 

of O3 properties such as temporal persistence and seasonal variability. Figure 3-3 shows 

nonurban, monthly average, daily maximum 8-h O3 concentrations in the five regions in the 

eastern United States from April to October 1993 to 2002. 

Regional differences are immediately apparent. Highest concentrations among all the 

regions are generally found in the Mid-Atlantic region (mean of 52 ppb), with highest values 

throughout the O3 concentration distribution except for the overall maximum. Lowest mean 

concentrations (42 ppb) are found in Florida. In the northern regions (the Northeast, Great 

Lakes) and the Mid-Atlantic region, highest median and peak concentrations are found in July, 

whereas in the Southwest region, highest median concentrations are found in August, with 

highest peaks in June and September, i.e., outside the warmest summer months. In Florida, 

highest monthly averaged median and peak concentrations are found during the spring. High O3 

concentrations tend to be most persistent (3 to 4 days of persistence) in the southern regions, less 

persistent in the Mid-Atlantic region (2 to 3 days) and least persistent in the northern regions 

(1 or 2 days). Such analyses could not be made for the western United States, in part because of 

the difficulty in finding regions with relatively coherent O3 properties as noted above for the 

eastern United States. 

Box plots showing the distributions of hourly average O3 concentrations for different types 

of rural sites for 2004 are given in Figures 3-4a (rural-agricultural), 3-4b (rural-forest), and 3-4c 

(rural-residential or commercial). Some associated metrics for vegetation exposures are given in 

Figures AX3-8 to AX3-10. Note that high O3 concentrations are found at sites that are classified 

as rural, such as Anne Arundel Co., MD; Yosemite NP, CA; and Crestline, CA. Land use 

designations do not usually give an accurate picture of exposure regimes in rural areas, because 

the land use characterization of “rural” does not imply that a specific location is isolated from 

3-6
Figure 3-3. Box plots showing daily maximum 8-h O3 averaged by month over 1993 to 

2002 in the five regions in the eastern United States derived by Lehman 

et al.(2004). The boxes define the interquartile range and the whiskers, the 

extreme values. 

Source: Lehman et al. (2004). 

3-7 
Figure 3-4a-c. Hourly average O3 concentrations observed at selected (a) ruralagricultural (b) rural-forested, and (c) rural-residential or commercial 

sites for 2004. The whiskers on the box plot represent the 10th and 

90th percentile concentrations. The “X”s above and below the whiskers 

are the values that fall below and above the 10th and 90th percentile 

concentrations. The dots inside the box represent the mean, for the 

statistic, at all sites. The number of observations is shown above each 

box plot. 

Source: Fitz-Simons et al. (2005). 

3-8 
anthropogenic influences. Rather, the characterization refers only to the current use of the land, 

not to the presence of sources. Since O3 produced from emissions in urban areas is transported 

to more rural downwind locations, elevated O3 concentrations can occur at considerable 

distances from urban centers. In addition, major sources of O3 precursors such as power plants 

and highways are located in nonurban areas and also produce O3 in these areas. Due to lower 

chemical scavenging in nonurban areas, O3 tends to persist longer in nonurban than in urban 

areas which tends to lead to higher exposures in nonurban areas influenced by anthropogenic 

precursor emissions. 

Ozone Air Quality Data at Relatively Remote Monitoring Sites (RRMS) 

Relatively Remote Monitoring Sites (RRMS) are sites located in national parks that tend to 

be less affected by obvious pollution sources than other sites. However, this does not mean that 

they are completely unaffected by local pollution, because of the large number of visitors to the 

national parks. 

Box plots showing the distribution of annual hourly averaged O3 concentrations at four 

RRMS are given in Figures 3-5a-d. It is important to characterize hourly average O3 

concentrations at RRMS so that assessments of the possible effects of O3 on human health and 

vegetation use concentration ranges that span the range of O3 concentrations found in the U.S. 

In many controlled exposure studies examining vegetation, O3 is filtered out of ambient air 

before it is admitted into the exposure chambers. As a result, O3 levels of only a few ppb are 

used as controls. 

As can be seen from Figures 3-5a-d, annual mean values of the daily maximum 8-h O3 

concentration have not changed much over the past 10 years of available data. Mean values 

typically range from about 0.020 ppm to about 0.040 ppm. Concentrations only rarely exceed 

0.080 ppm, in contrast to observations at other “rural” sites shown in Figures 3-4a-c. 

It is unlikely that distributions found at sites with low maximum hourly average 

concentrations in the western United States could represent those at sites in the eastern and 

midwestern United States because of regional differences in sources of precursors and transport 

patterns. Given the high density of sources in the eastern and midwestern United States, it is 

unclear whether a site could be found in either of these regions that would not be influenced by 

the transport of O3 from nearby urban areas. Thus, with the exception of the Voyageurs NP site 

3-9
Figure 3-5a-d. Daily 8-h maximum O3 concentrations observed at selected national park 

sites. The whiskers on the box plot represent the 10th and 90th percentile 

concentrations. The “X”s above and below the whiskers are the values that 

fall below and above the 10th and 90th percentile concentrations. The dots 

inside the box represent the mean. The number of observations is shown 

above each box plot. 

Source: Fitz-Simons et al. (2005). 

3-10 
in Minnesota, observations at RRMS are limited to those obtained in the western United States. 

However, not all national park sites in the West can be considered to be free of strong regional 

pollution influences, e.g., Yosemite NP (CA) as shown in Figure 3-4b. Maps showing the 

nationwide distribution of various metrics for vegetation exposures are given in Section AX3.2, 

Figures AX3-13 to AX3-27. 

3.3 SPATIAL VARIABILITY OF OZONE IN URBAN AREAS 

The spatial variability in O3 concentrations in 24 MSAs across the United States was 

examined. These MSAs were selected to provide (1) information helpful for risk assessments, 

(2) a general overview of the spatial variability of O3 in different regions of the country, and 

(3) insight into the spatial distribution of O3 in cities where health outcome studies have been 

conducted. Statistical analyses of the human health effects of airborne pollutants based on 

aggregate population time-series data have often relied on ambient concentrations of pollutants 

measured at one or more central sites in a given metropolitan area. In the particular case of 

ground-level O3 pollution, central-site monitoring has been justified as a regional measure of 

exposure mainly on the grounds that correlations between concentrations at neighboring sites 

measured over time are usually high. In MSAs with multiple monitoring sites, averages over the 

monitors have often been used to characterize population exposures. However, substantial 

differences in concentrations between monitors can exist even though concentrations measured 

at the monitoring sites are highly correlated, thus leading to the potential for exposure 

misclassification error. 

Metrics for characterizing spatial variability include the use of Pearson correlation 

coefficients (r), values of the 90th percentile (P90) of the absolute difference in O3 concentrations, 

and coefficients of divergence (COD)1

. These methods of analysis follow those used for 

characterizing PM2.5 and PM10-2.5 concentrations in Pinto et al. (2004) and in the latest edition of 

1 The COD is defined as follows: 

where xij and xik represent the 24-h average PM2.5 concentration for day i at site j and site k and p is the number of 

observations. 

3-11
the Particulate Matter Air Quality Criteria Document (PM AQCD) (U.S. Environmental Agency, 

2004a). However, the calculations were performed on an hourly basis rather than on a 24-h 

basis. Data were aggregated over local O3 seasons, the lengths of which vary from state to state. 

In several southwestern states, it lasts all year long. However, it typically last for 6 months 

in other areas, such as in New England, the mid-Atlantic states, the Midwest, and the Northwest 

(see Table AX3-1). 

P

Table 3-1 shows the chosen urban areas, the range of 24-h average O3 concentrations over 

the local O3 season from 1999 to 2001, the range of intersite correlation coefficients, the range of 

90 differences in O3 concentrations between site pairs, and the range in COD values. A COD of 

zero implies that values in both data sets are identical, and a COD of one indicates that the two 

data sets are uncorrelated, with no matching values from either data set. In general, statistics 

were calculated for partial MSAs. This was done so as to obtain reasonable lower estimates of 

the spatial variability that is present, as opposed to examining the consolidated MSAs. However, 

this could not be readily done for Boston, MA and New York, NY, so statistics were calculated 

for those consolidated MSAs. More detailed calculations for a subset of nine MSAs are given in 

Figures AX3-28 through AX3-36 in Section AX3.3. 

As can be seen from Table 3-1, no clearly discernible regional differences were found 

in the ranges of parameters analyzed. Additional urban areas would need to be examined to 

discern broadscale patterns. The data indicate considerable variability in the concentration 

fields. Mean O3 concentrations vary within individual urban areas by factors of 1.4 to 4 in 

Table 3-1. Intersite correlation coefficients show mixed patterns (i.e., in some urban areas all 

pairs of sites are moderately to highly correlated, while other areas show a larger range of 

correlations). As may be expected, those areas showing a smaller range of seasonal mean 

concentrations also show a smaller range of intersite correlation coefficients. However, there are 

a number of cases where sites in an urban area may be moderately to highly correlated, but show 

substantial differences in absolute concentrations. In many cases, P90 values can equal or exceed 

seasonal mean O3 concentrations. 

It is instructive to compare the metrics for spatial variability shown in Table 3-1 to those 

calculated for PM2.5 and PM10-2.5 in the PM AQCD (U.S. Environmental Agency, 2004a). The 

values for concentrations and concentration differences are unique to the individual species, but 

the intersite correlation coefficients and the COD values can be directly compared. In general, 

3-12
Table 3-1. Summary Statistics for the Spatial Variability of O3 (in ppm) in Selected Urban Areas in t

Urban Area 

Number of 

Sites 

Minimum 

Mean Conc. 

Maximum 

Mean Conc. 

Minimum 

Corr. Coeff. 

Maximum 

Corr. Coeff. 

Minimum 

P90 a 

Maximum 

P90 

M

Boston, MA 18 0.021 0.033 0.46 0.93 0.012 0.041 New York, NY 29 0.015 0.041 0.45 0.96 0.0080 0.044 Philadelphia, PA 12 0.020 0.041 0.79 0.95 0.011 0.036 Washington, DC 20 0.022 0.041 0.72 0.97 0.010 0.032 Charlotte, NC 8 0.031 0.043 0.48 0.95 0.012 0.038 Atlanta, GA 12 0.023 0.047 0.63 0.94 0.013 0.045 Tampa, FL 9 0.024 0.035 0.74 0.94 0.011 0.025 Detroit, MI 7 0.022 0.037 0.74 0.96 0.0090 0.027 Chicago, IL 24 0.015 0.039 0.38 0.96 0.0080 0.043 Milwaukee, WI 9 0.027 0.038 0.73 0.96 0.0090 0.025 St. Louis, MO 17 0.022 0.038 0.78 0.96 0.0090 0.031 Baton Rouge, LA 7 0.018 0.031 0.81 0.95 0.0090 0.029 Dallas, TX 10 0.028 0.043 0.67 0.95 0.011 0.033 Houston, TX 13 0.016 0.036 0.73 0.96 0.0090 0.027 Denver, CO 8 0.022 0.044 0.60 0.92 0.013 0.044 El Paso, TX 4 0.022 0.032 0.81 0.94 0.012 0.023 Salt Lake City, UT 8 0.029 0.048 0.52 0.92 0.012 0.043 Phoenix, AZ 15 0.021 0.058 0.29 0.95 0.011 0.057 Seattle, WA 5 0.015 0.038 0.63 0.94 0.0080 0.024 Portland, OR 5 0.015 0.036 0.73 0.91 0.011 0.025 Fresno, CA 6 0.030 0.047 0.90 0.97 0.0090 0.027 Bakersfield, CA 8 0.028 0.047 0.23 0.96 0.013 0.052 Los Angeles, CA 14 0.010 0.042 0.42 0.95 0.010 0.053 Riverside, CA 18 0.018 0.054 0.38 0.95 0.013 0.057 

3-13

a b P90 = 90th percentile absolute difference in concentrations. 

COD = coefficient of divergence for different site pairs. 
the variability in O3 concentrations is larger than for PM2.5 concentrations and comparable to that 

obtained for PM10-2.5. Intersite correlation coefficients in some areas (e.g., Philadelphia, PA; 

Atlanta, GA; Portland, OR) can be very similar for both PM2.5 and for O3. However, there is 

much greater variability in the concentration fields of O3 as evidenced by the much higher COD 

values. Indeed, COD values are higher for O3 than for PM2.5 in each of the urban areas 

examined. In all of the urban areas examined for O3, some site pairs are always very highly 

correlated with each other (i.e., r >0.9) as seen for PM2.5. These sites also show less variability 

in concentration and are probably influenced most strongly by regional production mechanisms. 

The above considerations indicate that caution should be observed in using data from the 

network of ambient O3 monitors to approximate community-scale human exposures. A similar 

conclusion was reached for PM using data from the PM2.5 FRM network, as indicated in 

Section 3.4 of the PM AQCD (U.S. Environmental Protection Agency, 2004a). 

3.3.1 Small-Scale Horizontal and Spatial Variability in 

Ozone Concentrations 

Ozone concentrations near roadways 

Apart from the larger scale variability in surface O3 concentrations, there is also significant 

variability on the micro-scale (< a few hundred meters), especially near roadways and other 

sources of emissions that react with O3. These sources are not confined to urban areas. Sources 

of emissions that react with O3 such as highways and power plants are also found in rural areas. 

Johnson (1995) described the results of studies examining O3 upwind and downwind of 

roadways in Cincinnati, OH. In these studies, O3 upwind of the roadway was about 50 ppb and 

values as high as this were not found again until distances of about 100 m downwind. The O3 

profile varied inversely with that of NO, as might be expected. For peak NO concentrations of 

30 ppb immediately downwind of the road, the O3 mixing ratio was about 36 ppb, or about 70% 

of the upwind value. The magnitude of the downwind depletion of O3 depends on the emissions 

of NO, the rate of mixing of NO from the roadway and ambient temperatures. So depletions 

of O3 downwind of roadways are expected, but with variable magnitude. In interpreting 

historical data, it should be noted that scavenging of O3 by NO near roadways was more 

pronounced before the implementation of stringent NOx emissions controls. 

3-14
Guidance for the placement of O3 monitors (U.S. Environmental Protection Agency, 1998) 

states a separation distance that depends on traffic counts. For example, a minimum separation 

distance of 100 m from a road with 70,000 vehicles per day (about 3,000 vehicles per hour) is 

recommended for siting an O3 monitor to avoid interference that would mean a site is no longer 

representative of the surrounding area. An average rate of about 3,000 vehicles per hour passing 

by a monitoring site implies a road with rather heavy traffic. As noted in Section AX3.3.1 for 

the Lakewood, CA monitoring, O3 levels are lower at sites located near traffic than those located 

some distance away, and the scavenging of O3 by emissions of NO from roadways is a major 

source of spatial variability in O3 concentrations. 

Vertical Variations in Ozone Concentrations 

In addition to horizontal variability in O3 concentrations, consideration must also be given 

to variations in the vertical profile of O3 in the lowest layers of the atmosphere. The planetary 

boundary layer (PBL) consists of an outer and an inner portion. The inner part extends from the 

surface to about one-tenth the height of the PBL. Winds and transported pollutants, such as O3, 

are especially susceptible to interactions with obstacles, such as buildings and trees in the inner 

boundary layer (atmospheric surface layer) (e.g., Garratt, 1992). Inlets to ambient monitors 

(typically at heights of 3 to 5 meters) are located in, and human and vegetation exposures occur 

in this part of the boundary layer. 

Photochemical production and destruction of O3 occur throughout the PBL. However, O3 

is also destroyed on the surfaces of buildings, vegetation, etc. In addition, O3 is scavenged by 

NO emitted by motor vehicles and soils. These losses imply that the vertical gradient of O3 

should always be directed downward. The magnitude of the gradient is determined by the 

intensity of turbulent mixing in the surface layer. 

Most work characterizing the vertical profile of O3 near the surface has been performed in 

nonurban areas with the aim of calculating fluxes of O3 and other pollutants through forest 

canopies and to crops and short vegetation, etc. Corresponding data are sparse for urban areas. 

However, monitoring sites are often set up in open areas such as parks and playgrounds where 

surface characteristics may resemble those in rural areas more than those in the surrounding 

urban area. The vertical profiles of O3 measured over low vegetation is shown in Figure 3-6. 

These measurements were obtained as part of a field campaign to measure the fluxes of several 

3-15
Figure 3-6. Vertical profile of O3 obtained over low vegetation. Values shown are relative 

to concentrations at 4 m above the surface. Ozone concentrations for stable 

and unstable conditions were 41.3 and 24.1 ppb, and average O3 concentration 

weighted by stability class was 33.1 ppb at 4 m. 

Source. Horváth et al. (1995). 

gas and aerosol phase pollutants using the gradient-flux technique (Horváth et al., 1995). The 

labels stable and unstable in the figure refer to atmospheric stability conditions and average 

represents the overall average. Ozone concentrations were normalized relative to their values at 

a height of 4 m. As can be seen from the figure, there was a decrease of about 20% in going 

from a height of 4 m down to 0.5 m above the surface during stable conditions, but O3 decreased 

by only about 7% during unstable conditions. The average decrease was about 10% for all 

measurements. As might be expected, O3 concentrations at all heights were very highly 

correlated with one another. Of course, these values represent averages and there is scatter about 

them. Under strongly stable conditions, they fall off toward the surface. However, these 

conditions tend to occur mainly during night and the stability regime during the day in urban 

areas tends more toward instability because of the urban heat island effect. Figure 3-7 shows the 

vertical profile of O3 measured in a spruce forest (Horváth et al., 2003). The fall off in O3 for 

3-16
Figure 3-7. Vertical profile of O3 obtained in a spruce forest. Values shown are relative to 

concentrations at 19 m above the surface. Mean tree height is 14.5 m. Ozone 

concentrations for stable and unstable conditions were 36.7 and 33.8 ppb, and 

the average O3 concentration weighted by stability class was 34.6 ppb at 19 m. 

Source: Horváth et al. (2003). 

this case is due to uptake by trees, reaction with ambient NO and with NO emitted by the soil 

in the forest, and reaction with hydrocarbons emitted by the trees in addition to deposition on 

the surface. 

3.4 DIURNAL AND SEASONAL VARIABILITY OF OZONE 

Diurnal Variability 

Diurnal variations in O3 at a given location are controlled by a number of factors, such as 

the relative importance of transport versus local photochemical production and loss rates, the 

timing for entrainment of air from the nocturnal residual boundary layer and the diurnal 

variability in mixing layer height. 

3-17
Diurnal Patterns in the Nationwide Data Set 

Composite urban, diurnal variations in hourly averaged O3 for April through October 2000 

to 2004 are shown in Figure 3-8. As can be seen from Figure 3-8, daily 1-h maxima tend to 

occur in mid-afternoon and daily 1-h minima tend to occur during the early morning. However, 

there is also considerable spread in these times. Therefore, some caution must be exercised in 

extrapolating results from one city to another and when attempting to judge the time of day when 

the daily 1-h maximum occurs. 

Figure 3-8. Composite, nationwide diurnal variability in hourly averaged O3 in 

urban areas. Values shown are averages from April to October 2000 to 

2004. Boxes define the interquartile range and the whiskers, the minima 

and maxima. 

Source: Fitz-Simons et al. (2005). 

Corresponding data for 8-h average O3 variations are shown in Figure 3-9. As can be seen 

from Figure 3-9, daily maximum eight hour O3 concentrations tend to occur from about 10 a.m. 

to about 6 p.m. As can be seen from Figure 3-9, they can also occur at slightly different times 

3-18
Figure 3-9. Composite, nationwide diurnal variability in 8-h average O3 in urban areas. 

Values shown are averages from April to October 2000 to 2004. Boxes define 

the interquartile range and the whiskers, the minima and maxima. The hour 

refers to the start of the 8-h averaging period. 

Source: Fitz-Simons et al. (2005). 

and the variation in the 8-h averages is smoother than for the 1-h averages. The minima in the 

8-h averages tend to occur starting at about midnight. 

Diurnal Patterns in EPA’s 12 Cities 

The diurnal variability of hourly averaged O3 in the twelve urban areas considered for 

inclusion in EPA’s human health exposure assessment risk assessment for the current review is 

illustrated in Figures 3-10a-l for April to October. Daily maximum 1-h concentrations tend to 

occur in mid-afternoon. However, as can be seen from the figures, the diurnal patterns vary 

from city to city, with high values ($0.100 ppm) also occurring either late in the evening as in 

Boston, past midnight as in Los Angeles and Sacramento, or midmorning as in Houston. 

Typically, high values such as these are found during the daylight hours in mid to late afternoon. 

The reasons for the behavior of O3 during the night at the above-mentioned locations are not 

3-19
Figure 3-10a-f. Diurnal variability in hourly averaged O3 in selected urban areas. Values 

shown are averages from April to October 2000 to 2004. Boxes define the 

interquartile range and the whiskers, the minima and maxima. 

Source: Fitz-Simons et al. (2005). 

3-20 
Figure 3-10g-l. Diurnal variability in hourly averaged O3 in selected urban areas. Values 

shown are averages from April to October 2000 to 2004. Boxes define the 

interquartile range and the whiskers, the minima and maxima. 

Source: Fitz-Simons et al. (2005). 

3-21 
clear. Measurement issues may be involved, or there may be physical causes such as transport 

phenomena, as discussed in Chapter 2. As discussed in Chapter 2, and in greater detail in 

Section AX2.3.3, nocturnal low level jets are capable of producing secondary O3 maxima 

at night. 

The diurnal variability of O3 averaged over 8 hours in the same twelve urban areas is 

shown in Figures 3-11a-l. The diurnal patterns of 1-h O3 averages are broadly similar to 8-h 

averages. Typically, although the 8-h daily maximum occurs between 10 a.m. and 6 p.m., actual 

starting and ending times can differ from these characteristic times depending on location. 

For example, as shown in Figures 3-11a for Boston and 3-11k for Sacramento, the highest 8-h 

daily maximum values can start in mid-afternoon and extend into late evening. These results 

suggest that transport processes are playing the dominant role in determining the timing of the 

highest daily maxima in these areas. 

On days with high 1-h daily maximum concentrations (e.g., $0.12 ppm), the maxima 

tend to occur in a smaller time window centered in the middle of the afternoon, compared to 

days on which the maximum is lower. For example, on high O3 days the 1-h maximum occurs 

from about 11 a.m. to about 6 p.m. However, on days for which the 1-h daily maximum 

is #0.080 ppm, the daily maximum can occur at any time during the day or night, with only a 

50% probability that it occurs between 1 and 3 p.m., in each of the 12 cities. (The time of day 

when the daily maximum 1-h O3 concentration occurs is illustrated for four of the cities in 

Figures AX3-45a-d.). Photochemical reactions in combination with diurnal emissions patterns 

are expected to produce mid-afternoon peaks in urban areas. These results suggest that transport 

from outside the urban airshed plays a major role in determining the timing of the daily maxima 

for low peak O3 levels. This pattern is typical for the Los Angeles-Long Beach-Riverside, CA 

area even on high O3 days. 

The same general timing patterns are found for 1-h daily maximum O3 concentrations as 

for the daily maximum 8-h average O3 concentration. As mentioned above, the daily maximum 

8-h O3 concentrations are generally found between the hours of 10 a.m. and 6 p.m. However, 

there are a significant number of days when this is not the case, e.g., for high values in Houston, 

TX and Los Angeles, CA, or in general for low values at any of the cities examined. (The time 

of day when the daily maximum 8-h average O3 concentrations occurs is shown for four cities in 

Figures AX3-46a-d.). Although the 8-h average O3 concentration is highly correlated with the 

3-22
Figure 3-11a-f. Diurnal variability in 8-h O3 in selected urban areas. Values shown 

are averages from April to October 2000 to 2004. Boxes define the 

interquartile range and the whiskers, the minima and maxima. 

The hour refers to the start of the 8-h averaging period. 

Source: Fitz-Simons et al. (2005). 

3-23 
Figure 3-11g-l. Diurnal variability in 8-h O3 in selected urban areas. Values shown 

are averages from April to October 2000 to 2004. Boxes define the 

interquartile range and the whiskers, the minima and maxima. 

The hour refers to the start of the 8-h averaging period. 

Source: Fitz-Simons et al. (2005). 

3-24 
daily maximum 1-h average O3 concentration, there are situations where the daily maximum 8-h 

average O3 concentration might be driven by very high values in the daily maximum 1-h average 

O3 concentration as illustrated in Figure 3-10j for Houston, TX. In cases such as these, the 

predicted 8-h average may overestimate the short-term O3 concentration later in the day. 

The patterns of diurnal variability for both 1-h and 8-h averages have remained quite stable 

over the 15-year period from 1990 to 2004, with times of occurrence of the daily maxima 

varying by no more than an hour from year to year in each of the 12 cities. 

Weekday/Weekend Differences 

Differences in the diurnal behavior of O3 have been observed in a number of cities (e.g., 

Heuss et al., 2003). Figures 3-12a-h show the contrast in the patterns of hourly averaged O3 in 

the greater Philadelphia, Atlanta, Houston and Los Angeles areas from weekdays to weekends. 

Daily maximum concentrations occur basically at the same time on either weekdays or 

weekends. Differences are apparent in the hourly concentrations, especially in the extreme 

values. Weekday/weekend differences in 8-h average O3 concentrations are shown in 

Figures 3-13a-h. As can be seen from a comparison of the weekend versus weekday patterns, 

there is a tendency for the lowest values in the distribution to be higher on weekends than on 

weekdays. Lower traffic volumes, in particular diesel truck traffic, lead to less NO emissions 

and titration of O3 on weekends. The spike in values for Houston in midmorning, shown in 

Figure 3-12f, resulted from the release of highly reactive hydrocarbons from the petrochemical 

industry (which could occur on any day of the week). Otherwise, the maximum O3 

concentrations could be seen to occur on the weekdays as they do in Philadelphia and Atlanta, 

in contrast to Los Angeles. Indeed, the diurnal pattern in Houston is similar to that observed in 

Atlanta on weekdays, indicating some overall similarity in the sources of precursors of O3. 

Spatial Variability in Diurnal Patterns in Urban Areas 

Daily maxima in either the 1-h or 8-h averages do not necessarily occur at the same time of 

day at each site in the 12 cities, and the diurnal pattern observed at individual sites can vary from 

the composites shown in Figures 3-8 and 3-9. Differences between sites are not only related to 

the distance between them; they also depend on nearby sources, such as highways, affecting one 

site to a greater extent than another. For example, in the Los Angeles basin, daily 1-h maxima 

3-25
Figure 3-12a-d. Diurnal variations in hourly averaged O3 on weekdays and weekends in 

four cities. Values shown represent averages from May to September 

of 2004. Boxes define the interquartile range and the whiskers, the 

minima and maxima. 

Source: Fitz-Simons et al. (2005). 

are reached in the late afternoon. However, in Riverside (which is typically located downwind 

of Los Angeles) there are sites in which the maximum is reached much earlier, because these 

sites are located close to highways. 

The general pattern that emerges from the site-to-site variability within the urban areas 

examined is that peaks in 1-h average concentrations are higher and tend to occur later at 

downwind sites than in the urban cores. To the extent that monitoring sites are either near to or 

3-26
Figure 3-12e-h. Diurnal variations in hourly averaged O3 on weekdays and weekends in 

four cities. Values shown represent averages from May to September 

of 2004. Boxes define the interquartile range and the whiskers, the 

minima and maxima. 

Source: Fitz-Simons et al. (2005). 

remote from sources of precursors in urban/suburban areas, the behavior of O3 will follow these 

basic patterns. Similar relations are found for the 8-h average O3 concentrations. Differences in 

diurnal patterns between sites in urban cores and sites downwind of urban cores are illustrated 

in Figures AX3-49a-b to AX3-51a-b for 1-h average O3 and in Figures AX3-52a-b to AX3-54a-b 

for Detroit, MI, St. Louis, MO, and Riverside, CA areas. 

3-27
Figure 3-13a-d. Diurnal variations in 8-h average O3 on weekdays and weekends in four 

cities. Values shown represent averages from May to September of 2004. 

Boxes define the interquartile range and the whiskers, the minima 

and maxima. The hour refers to the start of the 8-h averaging period. 

Source: Fitz-Simons et al. (2005). 

Seasonal Variability 

It should not be assumed that highest O3 levels are confined to the summer. Highest 

average O3 concentrations generally occur at background monitoring sites at midlatitudes in the 

Northern Hemisphere during late winter and spring versus summer, as for urban sites or for 

nonurban sites heavily affected by regional pollution sources. 

3-28
Figure 3-13e-h. Diurnal variations in 8-h average O3 on weekdays and weekends in four 

cities. Values shown represent averages from May to September of 2004. 

Boxes define the interquartile range and the whiskers, the minima 

and maxima. The hour refers to the start of the 8-h averaging period. 

Source: Fitz-Simons et al. (2005). 

High O3 values are also found at some of the 12 cities outside of summer. The seasonal 

behavior of O3 varies across the 12 cities. In most northern cities, the extreme values of the daily 

maximum 8-h average O3 concentration are a little more than half of those during the warm 

season, and the ratios of the medians are more similar as can be judged by comparison of Figures 

3-11a-l with Figures 3-14a-l. Differences are even smaller for the southern cities. Indeed, some 

of the highest O3 values are found in the Houston CSA outside of summer (Figure 3-14j). 

3-29
Figure 3-14a-f. Diurnal variability in 8-h average O3 in selected urban areas. Values 

shown are averages from November to March 2000 to 2004. Boxes define 

the interquartile range and the whiskers, the minima and maxima. The 

hour refers to the start of the 8-h averaging period. 

Source: Fitz-Simons et al. (2005). 

3-30 
Figure 3-14g-l. Diurnal variability in 8-h average O3 in selected urban areas. Values 

shown are averages from November to March 2000 to 2004. Boxes define 

the interquartile range and the whiskers, the minima and maxima. The 

hour refers to the start of the 8-h averaging period. 

Source: Fitz-Simons et al. (2005). 

3-31 
Diurnal Patterns in Nonurban Areas 

Composite diurnal patterns of O3 are shown in Figure 3-15 for hourly averaged O3 and in 

Figure 3-16 for 8-hour average O3 at rural (CASTNET) sites. As can be seen from a comparison 

of Figures 3-15 and 3-16 with Figures 3-8 and 3-9, diurnal patterns of O3 are smoother and 

shallower at the rural sites than at the urban sites. Maxima in hourly average O3 also tend to 

occur in afternoon. However, highest concentrations observed during any particular hour at 

night at the CASTNET sites (~0.130 ppm) are substantially higher than observed in urban areas 

(<0.100 ppm) and daily 1-h maxima at CASTNET sites have exceeded 0.150 ppm. The diurnal 

variations in 8-h average O3 concentrations are also much smaller at the CASTNET sites than at 

the urban sites. Note also that the maxima in 8-h average O3 concentrations are higher at the 

CASTNET sites than at the urban sites. 

3.5 TRENDS IN OZONE CONCENTRATIONS 

Year-to-year variability in the nationwide May to September, mean daily maximum 8-h O3 

concentrations are shown in Figure 3-17. The corresponding year-to-year variability in the 95th 

percentile concentrations is shown in Figure 3-18. Data flagged because of quality control issues 

were removed with concurrence of the local monitoring agency. Only days with data for 18 of 

24 hours were kept, and a minimum of 115 of 153 days were required in each year. Missing 

years were filled in using simple linear interpolation, as done in EPA Trends reports. Year-toyear variability in the 95th percentile values of the daily maximum 8-h O3 concentrations are 

shown in Figure 3-18. Sites considered in this analysis are shown in the map in Figure AX3-3. 

As was shown in Figures 3-1 and 3-2, most sites are located in the East. As can be seen from 

Figure 3-17, the highest O3 concentrations have tended to decrease over the past 15 years, while 

there has been little change in O3 concentrations near the center of the distribution. This is 

consistent with observations in Europe (Volz-Thomas et al., 2003). Mean O3 concentrations 

were slightly lower in 2003 and 2004 than in earlier years. The summer of 2003 was slightly 

cooler than normal in the East (Levinson and Waple, 2004) and the summer of 2004 was much 

cooler than normal in the East (Levinson, 2005) accounting in part for the dip in O3 during these 

two years. Observations of O3 at a number of sites in the Northern Hemisphere likewise do not 

show convincing evidence of strong upward trends during the 1990s (Oltmans et al., 1998). 

3-32
Figure 3-15. Composite diurnal variability in hourly O3 concentrations observed at 

CASTNET sites. Values shown are averages from April to October 2000 to 

2004. Boxes define the interquartile range and the whiskers, the minima 

and maxima. 

Source: Fitz-Simons et al. (2005). 

Figure 3-16. Composite diurnal variability in 8-h O3 concentrations observed at 

CASTNET sites. Values shown are averages from April to October 2000 to 

2004. Boxes define the interquartile range and the whiskers, the minima 

and maxima. The hour refers to the start of the 8-h averaging period. 

Source: Fitz-Simons et al. (2005). 

3-33 
Figure 3-17. Year-to-year variability in nationwide mean daily maximum 8-h O3 

concentrations. The whiskers on the box plot represent the 10th and 90th 

percentile concentrations. The “X”s above and below the whiskers are the 

values that fall below and above the 10th and 90th percentile concentrations. 

The dots inside the box represent the mean, for the statistic, at all sites. 

Source: Fitz-Simons et al. (2005) 

There may even have been a slight increase in O3 concentrations near the bottom of the 

distribution throughout the monitoring period. This would be consistent with data obtained in 

Europe, showing that O3 minima increased during the 1990s because of reduced titration of O3 

by reaction with NO in response to reductions in NOx emissions. As a result, the concentration 

of OX (NO2 + O3) shows little if any increase at all at the European sites (Volz-Thomas et al., 

2003). Trends in compliance metrics such as the fourth highest daily maximum 8-h O3 

concentration can be found in the U.S. EPA Trends reports (http://www.epa.gov/airtrends) and 

so are not repeated here. 

3-34
Figure 3-18. Year-to-year variability in nationwide 95th percentile value of the daily 

maximum 8-h O3 concentrations. The whiskers on the box plot represent the 

10th and 90th percentile values for the statistic. The “X”s above and below 

the whiskers are the values that fall below and above the 10th and 90th 

percentile values. The dots inside the box represent the mean, for the 

statistic, at all sites. 

Source: Fitz-Simons et al. (2005). 

Figures 3-19a-h show year-to-year variability in mean daily 8-h O3 concentrations 

observed at selected national park sites across the United States. Figures 3-20a-h show year-toyear variability in the 95th percentile value of daily maximum 8-h O3 concentrations at the same 

sites shown in Figures 3-19a-h. The same criteria used for calculating values in Figures 3-17 

and 3-18 were used for calculating the May to September seasonal averages for the national 

parks shown in Figures 3-19a-h and 3-20a-h. Sites at 22 national parks met these criteria, and 

data for all 22 sites are given in Appendix AX3 in Figures AX3-66a-v and AX3-67a-v. 

However, several monitoring sites were moved during the period from 1990 to 2004. Sites were 

moved at Acadia NP in 1996, Joshua Tree NP in 1993, Mammoth Cave NP in 1996, Voyageurs 

NP in 1996, and Yellowstone NP in 1996. These moves often resulted in offsets in O3 and so 

3-35
Figure 3-19a-h. Year-to-year variability in mean daily maximum 8-h O3 concentrations at 

selected national park (NP), national wildlife refuge (NWR), and national 

monument (NM) sites. 

Source: Fitz-Simons et al. (2005) 

3-36 
Figure 3-20a-h. Year-to-year variability in 95th percentile of daily maximum 8-h O3 

concentrations at selected national park (NP), national wildlife refuge 

(NWR), and national monument (NM) sites. 

Source: Fitz-Simons et al. (2005). 

3-37 
trends for these locations have not been calculated (cf., Section AX3.6, Table AX3-9). As noted 

in The Ozone Report—Measuring Progress through 2003 (U.S. Environmental Protection 

Agency, 2004b), O3 trends in national parks in the South and the East are similar to nearby urban 

areas and reflect the regional nature of O3 pollution. For example, O3 trends in Charleston, SC 

and Charlotte, NC track those in nearby Cowpens NP and Cape Romaine NP in South Carolina; 

O3 in Knoxville and Nashville, TN tracks O3 in Great Smoky NP; O3 in Philadelphia, PA and 

Baltimore, MD tracks Brigantine NP in New Jersey; and New York, NY and Hartford, CT track 

O3 in Cape Cod NS. The situation is not as clear in the West, where some national parks are 

affected by local pollution sources (e.g., Lassen Volcanic National Park and Yosemite National 

Park, CA) more than others. However, data obtained at these sites still provide valuable 

information about the variability in regional background concentrations, especially since the 

West has not been broken down into regions as has been done by Lehman et al. (2004) for the 

East as shown in Figure 3-3. Comparison of Figures 3-19a-h and 3-20a-h shows that O3 

concentrations near the center of the distribution do not necessarily track those at the upper end, 

as pointed out earlier for the nationwide composite data set. Trends in the 98th and 95th 

percentiles, and mean O3 concentrations for National Park sites presented in Table AX3-9 

demonstrate this point. In addition, Table AX3-9 shows that trends reversed direction in going 

from the 98th to 95th percentile values. 

Caution should be exercised in using trends calculated at national parks to infer 

contributions from distant sources either inside or outside of North America, because of the 

influence of regional pollution. For example, using a 15-year record of O3 from Lassen Volcanic 

NP and data from two aircraft campaigns, and observations spanning 18 years from five U.S. 

west coast marine boundary layer sites, Jaffe et al. (2003) have estimated that the amount of O3 

in air arriving from the Eastern Pacific in spring has increased by approximately 10 ppb from the 

mid-1980s to the present. They suggested this positive trend might be due to increases of 

emissions of O3 precursors in Asia. However, positive trends in O3 were found during all 

seasons. Although the Lassen Volcanic NP site is not close to any major emission sources or 

urban centers, maximum hourly average O3 concentrations of >0.080 ppm (during April-May) 

and >0.100 ppm (during the summer) occur at Lassen Volcanic NP, reflecting the influence of 

sources of O3 precursors in the area. Thus, although there is evidence that O3 levels may be 

3-38
increasing at some rural locations, there is also evidence that O3 levels at other locations have 

either not increased or have decreased over the same period. 

3.6 RELATIONSHIPS BETWEEN OZONE AND OTHER SPECIES 

Correlations between Ozone and Other Species 

In order to understand relationships among atmospheric species, an important distinction 

must be made between primary (directly emitted) species and secondary (photochemically 

produced) species. In general, it is likely that primary species will be highly correlated with 

other primary species, and that secondary species will be highly correlated with other secondary 

species. By contrast, primary species are less likely to be correlated with secondary species. 

Secondary reaction products tend to correlate with each other, but there is considerable variation. 

Some species (e.g., O3 and organic nitrates) are closely related photochemically and are highly 

correlated. Others (e.g., O3 and H2O2) show a more complex correlation pattern. Further details 

are given in Annex AX3 in Section AX3.7. 

Relationships between primary and secondary components are illustrated by considering 

data for O3 and PM2.5. Ozone and PM2.5 concentrations observed at a monitoring site in 

Fort Meade, MD are plotted as binned means for different intervals in Figure 3-21, based on data 

collected between July 1999 and July 2001. As can be seen from the figure, PM2.5 regarded as a 

function of O3 increases to the left of the inflection point (at about 30 ppbv O3) and also 

increases with O3 to the right of the inflection point. Data to the left of the minimum in PM2.5 

were collected mainly during the cooler months of the year, while data to the right of the 

minimum were collected during the warmer months. This situation arises because PM2.5 

contains a large secondary component during the summer and has a larger primary component 

during winter. During the winter, O3 comes mainly from the free troposphere, above the 

planetary boundary layer and, thus, may be considered a tracer for relatively clean air, and it is 

titrated by NO in the polluted boundary layer. Unfortunately, data for PM2.5 and O3 are collected 

concurrently at relatively few U.S. sites throughout an entire year. So these results, while highly 

instructive, are not readily extrapolated to areas where appreciable photochemical activity occurs 

throughout the year. Ito et al. (2005) examined the relation between PM10 and O3 on a seasonal 

basis in several urban areas (cf., Figure 7-24). Although PM10 contains proportionately more 

3-39
Figure 3-21. Binned mean PM2.5 concentrations versus binned mean O3 concentrations 

observed at Fort Meade, MD from July 1999 to July 2001. 

Source: Chen (2002). 

primary material than does PM2.5, relations similar to those shown in Figure 3-21 are found, 

reflecting the dominant contribution of PM2.5 to PM10. 

Other Oxidants 

Measurements of gas phase peroxides in the atmosphere were reviewed by Lee et al. 

(2000). Ground level measurements of H2O2 taken during the 1970s indicated values of 180 ppb 

in Riverside, CA and 10 to 20 ppb during smog episodes in Claremont and Riverside, with 

values approaching 100 ppb in forest fire plumes. However, later surface measurements always 

found much lower values. For example, in measurements made in Los Angeles and nearby areas 

in the 1980s, peak values were always less than about 2 ppb and in a methods intercomparison 

study in Research Triangle Park, NC in June 1986, concentrations were <2.5 ppbv. Higher 

values ranging up to 5 ppb were found in a few other studies in Kinterbish, Alabama and 

Meadview, Arizona. Several of these studies found strong diurnal variations (typically about a 

3-40
factor of three) with maximum values in the mid-afternoon and minimum values in the early 

morning. Mean concentrations of organic hydroperoxides at the surface at Niwot Ridge, CO in 

the summer of 1988 and State Park, GA during the summer of 1991 were all less than a few ppb. 

Aircraft measurements of hydroperoxide (H2O2, CH3OOH and HOCH2OOH) 

concentrations were made as part of the Southern Oxidants Study intensive campaign in 

Nashville, TN in July 1995 (Weinstein-Lloyd et al., 1998). The median concentration of total 

hydroperoxides in the boundary layer between 1100 and 1400 CDT was about 5 ppbv, with more 

than 50% contribution from organic hydroperoxides. Median O3 was about 70 ppbv at the same 

time. The concentrations of the hydroperoxides depended strongly on wind direction with values 

about 40% lower when winds originated from the N/NW as opposed to the S/SW suggesting that 

local source areas were important. 

Peroxyacetylnitrate (PAN) is produced during the photochemical oxidation of a wide range 

of VOCs in the presence of NOx. It is removed by thermal decomposition and also by uptake to 

vegetation (Sparks et al., 2003; Teklemariam and Sparks, 2004). PAN is the dominant member 

of the broader family of peroxyacylnitrates (PANs) which includes as other significant 

atmospheric components peroxypropionyl nitrate (PPN) of anthropogenic origin, and 

peroxymethacrylic nitrate (MPAN) produced from oxidation of isoprene. Measurements and 

models show that PAN in the United States includes major contributions from both 

anthropogenic and biogenic VOC precursors (Horowitz et al., 1998; Roberts et al., 1998). 

Measurements in Nashville during the 1999 summertime Southern Oxidants Study (SOS) 

showed PPN and MPAN amounting to 14% and 25% of PANs respectively (Roberts et al., 

2002). Measurements during the TexAQS 2000 study in Houston indicated PAN concentrations 

of up to 6.5 ppbv (Roberts et al., 2003). PAN measurements in southern California during the 

SCOS97-NARSTO study indicated peak concentrations of 5-10 ppbv, which can be contrasted to 

values of 60 to 70 ppbv measured back in 1960 (Grosjean, 2003). Vertical profiles measured 

from aircraft over the U.S. and off the Pacific coasts show PAN concentrations above the 

boundary layer of only a few hundred pptv, although there are significant enhancements 

associated with long-range transport of pollution plumes from Asia (Kotchenruther et al., 2001a; 

Roberts et al., 2004). Decomposition of this anthropogenic PAN as it subsides over North 

America can lead to significant O3 production, enhancing the O3 background (Kotchenruther 

et al., 2001b; Hudman et al., 2004). 

3-41
Oxidants are also present in airborne cloud droplets, rain drops and particulate matter. 

Measurements of hydroperoxides, summarized by Reeves (2003), are available mainly for 

hydrometeors, but are sparse for ambient particles. Venkatachari et al. (2005a) sampled the 

concentrations of total reactive oxygen species (ROS) in particles using a cascade impactor in 

Rubidoux, CA during July 2003. Although the species constituting ROS were not identified, the 

results were reported in terms of equivalent H2O2 concentrations. Unlike O3 and gas phase H2O2 

which show strong diurnal variability (i.e., about a factor of three variation between afternoon 

maximum and early morning minimum), the diurnal variation of particle phase ROS was found 

to be much weaker (i.e., less than about 20%) at least for the time between 8 a.m. and midnight. 

Because the ROS were measured in the fine aerosol size fraction, which has a lifetime with 

respect to deposition of much greater than a day, little loss is expected but their concentrations 

might also be expected to increase because of nighttime chemistry, perhaps involving NO3 

radicals. The ROS concentration, about 7 × 10!9 M/m3 (expressed as equivalent H2O2), was at 

most 1% that of O3 (6.2 to 38 × 10!7

 M/m3 or 15 to 90 ppb), with highest values at night. In a 

companion study conducted in Queens, NY during January and early February 2004, 

Venkatachari et al. (2005b) found much lower concentrations of ROS of about 1.2 × 10!9

 M/m3. 

However, O3 levels were also substantially lower, but ROS concentrations were still less than 

1% those of O3. It is of interest to note that gas phase OH concentrations measured at the same 

time ranged from about 7.5 × 104

/cm3 to about 1.8 × 106/cm3, implying the presence of 

significant photochemical activity even in New York City during winter. 

Co-Occurrence of Ozone with Other Pollutants 

The characterization of co-occurrence patterns under ambient conditions is important for 

relating human health and vegetation effects under ambient conditions to controlled research 

results as described in Annex AX3.8. Several attempts have been made to characterize gaseous 

air pollutant mixtures. The previous 1996 O3 AQCD discussed various patterns of pollutant 

mixtures of SO2, NO2, and O3. Pollutant combinations can occur at or above a threshold 

concentration at either the same or different times. 

The 1996 O3 AQCD noted that studies of the joint occurrence of gaseous NO2/O3 and 

SO2/O3 reached two conclusions: (1) hourly simultaneous and daily simultaneous-only 

co-occurrences are fairly rare (when both pollutants were present at an hourly average 

3-42
concentration $0.05 ppm) and (2) when co-occurrences are present, complex-sequential and 

sequential-only co-occurrence patterns predominate. Year-to-year variability was found to be 

insignificant. 

Using 2001 hourly data for O3 and NO2 and for O3 and SO2, the co-occurrence patterns for 

the data are similar to those of previous studies. As shown in Figure 3-22, fewer than 

10 co-occurrences of O3 and NO2 were found for most of the collocated monitoring sites. 

Likewise, Figure 3-23 shows that fewer than 10 co-occurrences of O3 and SO2 were found for 

most of the collocated monitoring sites analyzed. 

Figure 3-22. The co-occurrence pattern for O3 and nitrogen dioxide using 2001 data from 

the AQS. There is co-occurrence when hourly average concentrations of O3 

and another pollutant are both $0.05 ppm. 

Since 1999, monitoring stations across the United States have routinely measured 24-h 

average concentrations for PM2.5. Daily co-occurrence of PM2.5 and O3 over a 24-h period was 

also characterized. Because PM2.5 data are mostly summarized as 24-h average concentrations in 

the AQS database, a daily co-occurrence of O3 and PM2.5 was subjectively defined as an hourly 

average O3 concentration $0.05 ppm and a PM2.5 24-h concentration $40 μg/m3

 (corresponding 

to the EPA Air Quality Index, Level of Concern for PM2.5) occurring during the same 24-h 

3-43
Figure 3-23. The co-occurrence pattern for O3 and sulfur dioxide using 2001 data 

from AQS. There is co-occurrence when hourly average concentrations 

of O3 and another pollutant are both $0.05 ppm. 

period. Using 2001 data from the AQS database, the daily co-occurrence of PM2.5 and O3 was 

infrequent (Figure 3-24). Only limited data are available on the co-occurrence of O3 and other 

pollutants (e.g., acid precipitation and acidic cloudwater). In most cases, routine monitoring data 

are not available from which to draw general conclusions. 

3.7 POLICY RELEVANT BACKGROUND OZONE CONCENTRATIONS 

Background O3 concentrations used for purposes of informing decisions about NAAQS 

are referred to as Policy Relevant Background (PRB) O3 concentrations. Policy Relevant 

Background concentrations are those concentrations that would occur in the United States in the 

absence of anthropogenic emissions in continental North America (defined here as the United 

States, Canada, and Mexico). Policy Relevant Background concentrations include contributions 

from natural sources everywhere in the world and from anthropogenic sources outside these 

three countries. Background levels so defined facilitate separation of pollution levels that can be 

controlled by U.S. regulations (or through international agreements with neighboring countries) 

3-44
Figure 3-24. The co-occurrence pattern for O3 and PM2.5 using 2001 data from AQS. 

from levels that are generally uncontrollable by the United States. EPA assesses risks to human 

health and environmental effects from O3 levels in excess of PRB concentrations. Issues 

concerning the methodology for estimating PRB O3 concentrations are described in detail in 

Annex AX3, Section AX3.9. 

Contributions to PRB O3 include photochemical actions involving natural emissions of 

VOCs, NOx, and CO as well as the long-range transport of O3 and its precursors from outside 

North America and the stratospheric-tropospheric exchange (STE) of O3. Processes involved in 

STE are described in detail in Annex AX2.3. Natural sources of O3 precursors include biogenic 

emissions, wildfires, and lightning. Biogenic emissions from agricultural activities are not 

considered in the formation of PRB O3. 

Springtime maxima are observed at relatively remote (Annex AX3 and Figures 3-25a,b) 

national park sites, located mainly in the western United States and at a number of other 

relatively unpolluted monitoring sites throughout the Northern Hemisphere. The major issues 

concerning the calculation of PRB O3 center on the capability of the current generation of globalscale, three-dimensional, CTMs to simulate the causes of high O3 concentrations observed at 

monitoring sites in relatively unpolluted areas of the United States from late winter through 

spring (i.e., February through June). The issues raised do not affect interpretations of the causes 

3-45
Figure 3-25a. Monthly maximum hourly average O3 concentrations at Yellowstone

National Park (WY) in 1998, 1999, 2000, and 2001.

Source: U.S. Environmental Protection Agency (2003). 

Figure 3-25b. Hourly average O3 concentrations at Yellowstone National Park (WY) for 

the period January to December 2001. 

Source: U.S. Environmental Protection Agency (2003). 

3-46
of summertime O3 episodes as strongly. Summertime O3 episodes are mainly associated with 

slow- moving high-pressure systems characterized by limited mixing between the planetary 

boundary layer and the free troposphere, as noted in Annex AX2, Section AX2.3. 

A large number of case studies document the occurrence of STE mainly during winter and 

spring in mid- and high-latitudes in Europe, Asia, and North America. These studies were based 

on aircraft, satellite, and ground-based measurements. Considerable uncertainty exists in the 

magnitude of the exchange; however, these studies have found that STE occurs throughout the 

year, but with a distinct preference for the transport of O3 directly to the middle and lower 

troposphere during late winter and spring. Transport to the upper troposphere occurs throughout 

the year. 

Springtime maxima in tropospheric O3 observed at high latitudes are also associated with 

the winter buildup of O3 precursors and thermally labile reservoir species, such as PAN and 

other reactive nitrogen species. These pollutants originate from all continents in the Northern 

Hemisphere. Ozone precursor concentrations reach a maximum in late March; and as sunlight 

returns to the Arctic, photochemical reactions generate tropospheric O3 (Section AX3.9.1). 

The contribution of Asian sources to the U.S. levels is also largest during spring, reflecting the 

efficient lifting of Asian pollution ahead of cold fronts originating in Siberia and transport by 

strong westerly winds across the Pacific (e.g., Hudman et al., 2004). The longer lifetime of O3 

during spring also contributes to springtime maxima (Wang et al., 1998). 

Estimates of PRB concentrations cannot be obtained solely by examining measurements 

of O3 obtained at RRMS in the United States (Annex AX3, Section AX3.2.3) because of the 

long-range transport from anthropogenic source regions within North America. It should also be 

noted that it is impossible to determine sources of O3 without ancillary data that could be used as 

tracers of sources or to calculate photochemical production and loss rates. The current definition 

of PRB implies that only CTMs can be used to estimate the range of PRB values. On the 

synoptic and larger spatial scales at least, all evidence indicates that global CTMs are adequate 

tools to investigate the factors controlling tropospheric O3; and three-dimensional CTMs, as 

typified by Fiore et al. (2003) appear to offer the best methodology for estimating PRB 

concentrations that cannot be measured directly (Annex AX3, Section AX3.9.2), at least for 

averaging periods of longer than one hour. 

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Previous estimates of background O3 concentrations, based on different concepts of 

background, are given in Table 3-2. Results from global three-dimensional CTMs, where the 

background is estimated by zeroing anthropogenic emissions in North America (Table 3-8) are 

on the low end of the 25 to 45 ppbv range. Lefohn et al. (2001) have argued that frequent 

occurrences of O3 concentrations above 50 to 60 ppbv at remote northern U.S. sites in spring are 

mainly stratospheric in origin. Fiore et al. (2003) used a global CTM to determine the origin of 

the high-O3 events reported by Lefohn et al. (2001), and to conduct a more general quantitative 

analysis of background O3 as a function of season, altitude, and local O3 concentration. 

Table 3-2. Previous Estimates of Background O3 in Surface Air Over the United States 

Background 

Study Method Time Period Region Estimate (ppbv) 

Trainer et al. 

(1993) 

y-intercept of O3 vs. NOy-NOx 

regression linea 

Summer 1988 Eastern United 

States 

30-40b 

Hirsch et al. 

(1996) 

y-intercept of O3 vs. NOy-NOx 

regression line 

May-Sep 

1990-1994 

Harvard Forestc 25 (Sept) – 40 (May)d 

Altshuller 

and Lefohn 

y-intercept of O3 vs. NOy

regression line, and observations 

Apr-Oct 

1988-1993 

Continental 

United States 

25-45 (inland)e 

25-35 (coastal) 

(1996) at remote/rural sites 

Liang et al. Sensitivity simulation in a 3-D Full year Continental 20-30 (East)f 

(1998) model with anthropogenic NOx United States 20-40 (West) 

emissions in the continental U.S. (spring maximum) 

set to zero 

Lin et al. Median O3 values for the lowest 1990-1998 Harvard Forest 35 (fall) – 45 (spring)g 

(2000) 25th percentiles of CO and NOy

concentrations 

Fiore et al. O3 produced outside of the North Summer 1995 Continental 15-30 (East)h 

(2002) American boundary layer in a United States 25-35 (West) 

global 3-D model 

a NOy is the chemical family including NOx and its oxidation products; NOy-NOx denotes the chemically

 processed component of NOy. b 1300-1700 local time (LT) in flatland and valley sites; all daytime measurements at elevated sites. c rural site in central Massachusetts. 

d 1100-1700 EST hourly means. 

e seasonal 7-h (0900-1559) daylight average.

f 1300-1600 LT monthly mean. 

g daily max 8-h averages. 

h 1300-1700 average. 

Source: Fiore et al. (2003). 

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Figure 3-26 shows a comparison between observations obtained at CASTNet sites and 

model results of Fiore et al. (2003). They classified the CASTNet monitoring sites into 

low-lying sites (generally <1.5 km) and elevated sites (>1.5 km). All elevated sites are in the 

West. Results were then aggregated to construct the cumulative probability distributions shown 

in Figure 3-26 for the 58 low-altitude sites and the 12 high-altitude sites as well as for the three 

seasons. The calculated mean background at the surface sites in spring is 27 ppbv, compared to 

23 ppbv in summer and fall. At these sites, the background is highest for O3 concentrations near 

the center of the distribution, and it declines as total surface O3 concentrations increase, for 

reasons summarized below and discussed by Fiore et al. (2002). The observed O3 concentration 

thus serves a surrogate for meteorological variability (i.e., stagnant versus ventilated conditions), 

such that the background O3 is smaller on days when total O3 is highest. At the elevated sites, 

the calculated mean background is 36 ppbv in spring versus 30 ppbv in the summer and fall. 

Background concentrations in the fall resemble those in summer but show less variability and do 

not exceed 40 ppbv anywhere in this analysis. 

Major conclusions from the Fiore et al. (2003) study (discussed in detail in Annex AX3, 

Sections AX3.9.3 and AX3.9.4) are: 

• PRB O3 concentrations in U.S. surface air from 1300 to 1700 local time are generally 

15 to 35 ppbv. They decline from spring to summer and are generally <25 ppbv under 

the conditions conducive to high-O3 episodes. 

• PRB O3 concentrations can be represented as a function of season, altitude, and total 

surface O3 concentration, as illustrated in Figure 3-26. 

• High PRB concentrations (40 to 50 ppbv) occur occasionally at high-elevation sites 

(>1.5 km) in spring due to the free-tropospheric influence, including a 4- to 12-ppbv 

contribution from hemispheric pollution (O3 produced from anthropogenic emissions 

outside North America). These sites cannot be viewed as representative of low-elevation 

surface sites (Cooper and Moody, 2000), where the background is lower when O3 

>60 ppbv. 

• The stratospheric contribution to surface O3 is of minor importance, typically well

<20 ppbv. While stratospheric intrusions might occasionally elevate surface O3

at high-altitude sites, these events are rare.

Appropriate background concentrations should thus be allowed to vary as a function of 

season, altitude, and total O3 level. The diamonds in Figure 3-26 can be applied for this purpose. 

In particular, the depletion of the background during high-O3 events should be taken into account 

3-49
Figure 3-26. Estimates of background contribution to surface afternoon (13 to 17 LT) 

O3 concentrations in the United States as a function of local O3 concentration, 

site altitude, and season. The figure shows cumulative probability 

distributions of O3 concentrations for the observations (asterisks) and the 

model (triangles). The corresponding distribution of background O3 

concentrations is shown as grey diamonds. 

Source: Fiore et al. (2003). 

3-50 
(i.e., background O3 is depleted by reactions in the atmosphere and by deposition to the surface 

but is not replenished at a significant rate in the stable, polluted boundary layer). This depletion 

is shown in the right-hand panels of Figure 3-26 for the highest O3 values. Note that the model 

is generally able to reproduce the overall frequency distributions in Figure 3-26. Typically, 

models produce distributions flatter than are observed. Underpredictions, especially at the upper 

end of the frequency distribution during the warmer months, are likely related to sub-grid-scale 

processes that the model cannot resolve explicitly. The highest observed O3 concentrations in all 

three seasons and at all altitudes are associated with regional pollution (i.e., North American 

anthropogenic emissions), rather than stratospheric influence. 

PRB ozone is not a directly observable quantity and must therefore be estimated from 

models. Simple modeling approaches, such as the use of back-trajectories at remote U.S. sites to 

identify background conditions, are subject to errors involving the reliability of the trajectories, 

chemical production along the trajectories, and the hemispheric-scale contribution of North 

American sources to ozone in air masses originating outside the continent. They also cannot 

describe the geographical variability of the ozone background or the depletion of this 

background during pollution episodes. Global 3-D chemical transport models such as 

GEOS-Chem can provide physically-based estimates of the PRB and its variability through 

sensitivity simulations with North American anthropogenic sources shut off. These models are 

also subject to errors in the simulation of transport and chemistry, but the wealth of data that 

they provide on ozone and its precursors for the present-day atmosphere enables extensive 

testing with observations, and thus objective estimate of the errors on the PRB ozone values. 

We present here such an error analysis for the Fiore et al. (2003a) PRB ozone estimates, 

focusing on evaluation of their GEOS-Chem simulation for present-day conditions with ozone 

observations of particular relevance to the PRB. Comparisons to ozonesonde observations of the 

vertical distribution of tropospheric ozone at northern mid-latitudes are of particular interest. 

These have been documented in a number of GEOS-Chem publications (Bey et al., 2001; 

Liu et al., 2002; Martin et al., 2002; Fusco and Logan, 2003), including specifically over 

North America (Li et al., 2002a, 2005). Results indicate no significant bias, and agreement to 

generally within 5 ppbv for monthly mean concentrations at different altitudes. 

Fiore et al. (2002, 2003a,b) presented detailed evaluations of GEOS-Chem with ozone 

observations at U.S. surface sites. These evaluations focused on the afternoon hours (13-17 local 

3-51
time), when surface measurements are representative of a deep mixed layer that can be resolved 

with the model. At night, surface ozone depletion often takes place by titration or deposition 

under locally stratified conditions, but such conditions cannot be simulated with confidence by a 

global model. The issue is not only one of vertical resolution (the lowest layers in GEOS-Chem 

extend to 20, 50, 100, 200, and 400 m above the local surface) but also of horizontal resolution 

(2°x2.5°). The PRB estimates presented by Fiore et al. (2003a) are also for 13-17 local time, and 

lower values would apply at night. 

The GEOS-Chem evaluation of U.S. surface ozone presented by Fiore et al. (2003a) 

included monthly means, probability distributions, and time series at the sites previously used by 

Lefohn et al. (2001) to estimate background ozone. The simulated monthly mean concentrations 

in different seasons are typically within 5 ppbv of observations, with no significant bias, except 

in the southeast in summer when the model is 8-12 ppbv too high due to excessive background 

ozone advected in from the Gulf of Mexico. The time series comparisons for specific sites show 

that the model simulates the day-to-day variability of ozone and reveals no further bias. 

In particular, the model can reproduce the occurrences of relatively high ozone at remote sites 

previously reported by Lefohn et al. (2001), and shows that these can generally be explained by 

North American pollution. 

Goldstein et al. (2004) presented comparisons of GEOS-Chem and MOZART global 

model results with observations at Trinidad Head, California, during April-May 2002. The 

observations, filtered to remove local influence, averaged 41 ± 5 ppbv, as compared to 

GEOS-Chem (39 ± 5 ppbv) and MOZART (37 ± 9 ppbv). Neither model was successful at 

reproducing the weak day-to-day structure in the observations, but they showed no bias in the 

simulation of occasional >50 ppbv days. Hudman et al. (2004) found that GEOS-Chem 

underestimated ozonesonde measurements in the free troposphere over Trinidad Head during the 

same period by 10 ppbv on average, apparently due to insufficient regional stratospheric 

influence. This bias does not extend to the surface, likely because stratospheric influence greatly 

diminishes as free tropospheric air is diluted in the boundary layer. 

Several other papers have evaluated the GEOS-Chem simulation for surface ozone and its 

precursors over the United States. Fiore et al. (2003b) found a mean model bias of 6 ± 6 ppbv in 

surface air over the eastern United States in summer, reflecting the overestimate in the southeast; 

but they found that the model captured the large-scale spatial structure of the ozone observations 

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as revealed by an EOF analysis. The model also reproduces without significant bias the 

correlations of ozone with CO and NOy over North America and downwind (Fiore et al., 2002; 

Li et al., 2002b, 2004). Comparisons to satellite observations of NO2 and formaldehyde 

tropospheric columns over North America show excellent correspondence in the spatial patterns 

and no significant biases. 

In conclusion, we estimate that the PRB ozone values reported by Fiore et al. (2003a) for 

afternoon surface air over the United States are likely 10 ppbv too high in the southeast in 

summer, and accurate within 5 ppbv in other regions and seasons. The PRB ozone is likely 

lower at night than in the afternoon though Fiore et al. (2003a) do not quantify the magnitude of 

this diurnal effect. 

Chemistry transport models should be evaluated with observations given earlier in 

Chapter 3, in Annex AX3, and to simulate the processes causing the intra-day variability in O3 

concentrations shown in Figure 3-27 in addition to those summarized in Chapter 2. The diurnal 

patterns shown in Figure 3-27 do not fit the smooth pattern shown in Figure 3-15 and indicate 

processes capable of producing rapid rises in O3 at times when substantial photochemical 

activity is not present and may indicate stratospheric effects. Higher resolution models capable 

of spatially and temporally resolving stratospheric intrusions and capable of resolving O3 

variability on hourly timescales have not been applied to this problem. Ebel et al. (1991) have 

demonstrated that regional-scale CTMs could be used to study individual stratospheric 

intrusions. As an example of the utility of different types of models, Zanis et al. (2003) were 

able to forecast, observe, and model a stratospheric intrusion (maximum penetration depth was 

to slightly >2 km altitude) that occurred from June 20 to 21, 2001, over a large swath of central 

Europe. Roelofs et al. (2003) compared results from six global tropospheric CTMs with lidar 

observations obtained during that event and concluded that the models qualitatively captured the 

features of this intrusion. It was also found that the coarser resolution models overestimated 

transport to lower altitudes. The use of higher resolution models, perhaps nested inside the 

coarser resolution models, may have helped solve this problem. They would also better address 

issues related to temporal (i.e., 1-h versus 8-h averages) and spatial (i.e., populated versus 

remote areas) scales needed by policymakers. 

Although many of the features of the day-to-day variability of O3 at RRMS in the United 

States are simulated reasonably well by Fiore et al. (2003), uncertainties in the calculation of the 

3-53
Figure 3-27. Time-series of hourly average O3 concentrations observed at five national 

parks: Denali (AK), Voyageur (MN), Olympic (WA), Glacier (MT), and 

Yellowstone (WY). 

3-54 
temporal variability of O3 originating from different sources on shorter time scales must be 

recognized. The uncertainties stem in part from an underestimate in the seasonal variability in 

the STE of O3 (Fusco and Logan, 2003), the geographical variability of this exchange, and the 

variability in the exchange between the free troposphere and the planetary boundary layer in the 

model. 

Ideally, the predictions resulting from an ensemble of models should be compared with 

each other and with observations, so that the range of uncertainty inherent in the model 

predictions can be evaluated. In this regard, it should be noted that only one model (GEOSChem) is documented in the literature for estimating PRB concentrations. 

3.8 OZONE EXPOSURE IN VARIOUS MICROENVIRONMENTS 

Humans are exposed to O3 and related photochemical oxidants through the exchange 

boundary, the skin and the openings into the body such as the mouth, the nostrils, and punctures 

and lesions in the skin (U.S. Environmental Protection Agency, 1992; Federal Register, 1986). 

Inhalation exposure to O3 and related photochemical oxidants is determined by pollutant 

concentrations measured in the breathing zone that is not affected by exhaled air as the 

individual moves through time and space. A discussion of the basic terminology associated with 

exposure appears in AX3. 

Quantification of Exposure 

Ambient O3 concentrations vary with time of day (peaking during the latter portion of the 

day) and season and among locations. Consequently, exposure to O3 will change as a function of 

time of day and as an individual moves among locations. A hypothetical exposure is 

demonstrated in Figure 3-28. The actual dose received also changes during the day and is 

dependent on the O3 concentration in the breathing zone and the individual’s breathing rate, 

which is, in turn, dependent on the individual’s level of exertion. 

When measuring or modeling exposure to O3 and related photochemical oxidants 

consideration should be given to the diurnal weekly (weekday-weekend) and seasonal 

variability. Peak concentrations lasting for several hours typically occur toward the latter 

portion of the day during the summer months. Regional O3 episodes often co-occur with high 

3-55
Figure 3-28. Hypothetical exposure time profile: pollutant exposure as a function of time 

showing how the average exposure, integrated exposure, and peak exposure 

relate to the instantaneous exposure. (t2 – t1 = T) 

Source: U.S. Environmental Protection Agency (2004a). 

concentrations of airborne fine particles, making it difficult to assess O3 dynamics and exposure 

patterns. Also, while there are few indoor O3 sources, O3 will react with materials and other 

pollutants in the indoor environment in an analogous fashion to that occurring in the ambient 

atmosphere, potentially exposing subjects to other more toxic pollutants (Nazaroff and Weschler, 

2004; Lee and Hogsett, 1999; Wainman et al., 2000; Weschler and Shields, 1997). (See 

discussion on O3 chemistry and indoor sources and concentrations later in this chapter.). 

Personal Exposure and Ambient Concentrations 

The two approaches for measuring personal exposure are (a) the direct approach, using a 

personal exposure monitor (PEM) consisting of a passive sampler worn around the breathing 

zone, and (b) the indirect approach, which measures or estimates the O3 concentrations through 

the use of models or biomarkers. Both approaches are associated with measurement error. 

3-56
Although it is difficult to develop passive monitors for personal exposure measurements 

because of problems in identifying chemical or trapping reagents that can react with O3, several 

modified passive samplers have been developed for use in personal O3 exposure measurements 

(Bernard et al., 1999; Koutrakis et al., 1993; Avol et al., 1998b; Geyh et al., 1997, 1999). 

Personal exposure measurements using passive samplers show O3 exposures below those O3 

concentrations measured at outdoor stationary sites (Delfino et al., 1996; Avol et al., 1998a; 

Sarnat et al., 2000; Geyh et al., 2000). However, there is a strong correlation between O3 

measured at stationary sites and personal monitored concentrations (Liard et al., 1999; Brauer 

and Brook, 1997; Linn et al., 1996; Lee et al., 2004; Avol et al., 1998a; O’Neill et al., 2003) 

when the time spent outdoors, age, gender, and occupation of the subjects were considered. 

The indirect approach determines and measures the concentrations in all of the locations or 

“microenvironments.” The concept of microenvironments is important in the understanding of 

human exposure modeling. Often identified with a perfectly mixed compartment, 

microenvironments are more recently viewed as a controlled volume, indoors or outdoors, that 

can be characterized using a set of either mechanistic or phenomenological governing equations. 

This allows for a nonhomogeneous environment, including sources and sinks within the 

microenvironment. Microenvironments include indoor residences, other indoor locations, 

outdoors near roadways, other outdoor locations, and areas within vehicles. 

Microenvironmental Concentration and Ozone Exposure Models 

Outdoor concentrations of O3 are estimated either through emissions-based mechanistic 

modeling, or through ambient-data-based modeling. Emissions-based models determine the 

spatiotemporal fields of the O3 concentrations using precursor emissions and meteorological 

conditions as inputs. (They are described in Annex AX2.). The ambient-data-based models 

determine spatial or spatiotemporal distributions of O3 through the use of interpolation schemes. 

The kriging approach provides standard procedures for generating an interpolated O3 spatial 

distribution for a given period of time (Georgopoulos et al., 1997a,b). The Spatio-Temporal 

Random Field (STRF) approach has been used to interpolate monitoring data in both space and 

time (Christakos and Vyas, 1998a,b). The STRF approach can analyze information on temporal 

trends which cannot be directly incorporated by kriging. 

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Several approaches are available for modeling microenvironmental concentrations: 

empirical, mass balance, and detailed computational fluid dynamics (CFD) models. Empirical 

relationships provide the basis for future, “prognostic” population exposure models. Mass 

balance modeling is the most common approach used to model pollutant concentrations in 

enclosed microenvironments. Mass balance modeling ranges from very simple formulations, 

assuming ideal (homogeneous) mixing and only linear physicochemical transformations with 

sources and sinks, to models that account for complex multiphase chemical and physical 

interactions and nonidealities in mixing. Mass balance models take into account the effects of 

ventilation, filtration, heterogeneous removal, and direct emission as well as photolytic, thermal, 

and chemical reactions. The simplest form of the model is represented by the following 

differential equation: 

where dCIN is the indoor pollutant concentration (mass/volume), dt is time in hours, v is the air 

exchange rate, COUT is the outdoor pollutant concentration (mass/volume), V is the volume of the 

microenvironment, and S is the indoor source emission rate. When the model was used to 

estimate indoor O3 concentrations, indoor concentrations were found to be 33% of outdoor O3 

concentrations (Freijer and Bloemen, 2000). A more in-depth discussion of the mass balance 

model has been reported in Nazaroff and Cass (1986). The pNEM/O3 model, discussed later in 

this chapter, includes a sophisticated mass balance model for indoor and vehicle 

microenvironments (Johnson, 2003). CFD models take into account the complex, multiphase 

processes that affect indoor concentrations of interacting gas phase pollutants, such as the 

interactions of O3 with indoor sinks and sources (surfaces, gas releases) and with entrained gas 

(Sarwar et al., 2001, 2002; Sørensen and Weschler, 2002). 

Exposure modeling is often used in evaluating exposure to large populations over time. 

The use of models is complicated by the fact that O3 is a secondary pollutant with complex 

nonlinear and multiscale dynamics in space and time. Ozone is formed in the atmosphere 

through a series of chemical reactions involving precursor VOCs and NOx. Therefore, O3 

exposures may be affected by: (1) emission levels and spatiotemporal patterns of VOCs 

and NOx; (2) ambient atmospheric as well as indoor microenvironmental transport, removal 

3-58
and mixing processes; and (3) chemical transformations that take place over a multitude of 

spatial scales. The transformations are dependent on the presence of co-occurring pollutants and 

the nature of surfaces interacting with the pollutants. 

Exposure models may be classified as (1) potential exposure models, typically the 

maximum outdoor concentrations versus “actual” exposure, including locally modified 

microenvironmental outdoor and indoor exposures; (2) population versus “specific individual”-

based exposure models; (3) deterministic versus probabilistic models; and (4) observation versus 

mechanistic air quality model-driven estimates of spatially and temporally varying O3 

concentrations. 

There are several steps involved in defining exposure models. The steps are based on 

frameworks described in the literature over the last 20 years and the structure of various existing 

inhalation exposure models (NEM/pNEM, MENTOR/SHEDS, REHEX, TRIM.Expo also known 

as APEX, AIRPEX, AIRQUIS). The steps include (1) estimation/ determination of the 

background or ambient levels of O3; (2) estimation/determination of levels and temporal profiles 

of O3 in various microenvironments; (3) characterization of relevant attributes of individuals or 

populations under study (age, gender, weight, occupation, other physiological characteristics); 

(4) development of activity event or exposure event sequences; (5) determination of appropriate 

inhalation rates during the exposure events; (6) determination of dose; (7) determination of 

event-specific exposure and intake dose distributions for selected time periods; and 

(8) extrapolation of population sample (or cohort) exposures and doses to the entire populations 

of interest. Figure 3-29 provides a conceptual overview of a current exposure model. A more 

detailed overview of an exposure model can be found in Annex AX3. 

To estimate the actual O3 dose delivered to the lung, information on the concentration, 

minute ventilation rate, activity level, and the morphology of the respiratory tract are needed. 

Limited data have been compiled for ventilation rates for different age groups, both healthy and 

compromised individuals, at varies levels of activity (Klepeis et al., 1996, 2001; Avol et al., 

1998a; Adams, 1993). Based on the available information, the highest level of outdoor activity 

occurs during the spring and summer months, during the mid- to late afternoon and early 

evening—the times when O3 concentrations are highest. Children are likely more susceptible to 

the effects of O3 than other groups. School-age children spend more time outdoors engaged in 

high-level activities than do other groups and breath more air in than adults relative to body 

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Figure 3-29. Conceptual overview of an exposure model. Model inputs (e.g., activity 

patterns, ambient monitoring data, air exchange rates) are in round-corner 

boxes and model calculations are shown in rectangles. 

Source: Johnson et al. (1999). 

surface area, breathing frequency, and heart rate. Asthmatic children spend the same amount of 

time outdoors as other more healthy children but the time spent engaged in high levels of activity 

are less. 

Estimates of activity level have been compiled based on questionnaire data. The National 

Human Activity Pattern Survey (NHAPS), a probability-based telephone survey, was conducted 

in the early 1990s. The survey concluded that outdoor work-related activities were highest 

during the springtime and were more frequent during the morning and early afternoon. 

Exercise/sports-related activities were highest from noon to 3 p.m. during the summer months. 

During the spring months, exercise/sports-related activities were highest from mid- to late 

afternoon (Klepeis et al., 1996, 2001). A pilot study by Gonzales et al. (2003) evaluated the use 

of retrospective questionnaires for reconstructing past time-activity and location pattern 

information. Ozone concentration estimates using ambient stationary monitors and estimates 

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derived from diaries and questionnaires differed slightly. However, both estimates were greater 

than O3 personal exposure measurements. 

Existing comprehensive inhalation exposure models (NEM and pNEM) (Johnson, 2003; 

Johnson et al., 1996a,b,c, 1997), (MENTOR/SHEDS) Burke et al., 2001; McCurdy et al., 2000), 

and the APEX model treat human activity patterns as sequences of exposure events in which 

each event is defined by a geographic location and microenvironment and then assigned activity 

diary records from the CHAD (Consolidated Human Activities Database; 

www.epa.gov/chadnet1) (Glen et al., 1997; McCurdy, 2000; McCurdy et al., 2000). There are 

now about 22,600 person-days of sequential daily activity pattern data in CHAD representing all 

ages and both genders. The data for each subject consist of one or more days of sequential 

activities, in which each activity is defined by start time, duration, activity type (140 categories), 

and microenvironment classification (110 categories). Activities vary from 1 min to 1 h in 

duration. Activities longer than 1 h are subdivided into clock-hour durations to facilitate 

exposure modeling. A distribution of values for the ratio of oxygen uptake rate to body mass 

(referred to as metabolic equivalents or METs) is provided for each activity type listed. A table 

listing the activity patterns included in CHAD appears in AX3. 

pNEM divides the population of interest into representative cohorts based on the 

combinations of demographic characteristics (age, gender, and employment), home/work 

district, and residential cooking fuel. APEX and MENTOR/SHEDS generate a population 

demographic file containing a user-defined number of person-records for each census tract of 

the population based on proportions of characteristic variables (age, gender, employment, 

and housing) obtained for the population of interest, and then assigns the matching activity 

information from CHAD to each individual record of the population based on the characteristic 

variables. 

The APEX model is capable of simulating individual movement through time and space to 

provide estimates of exposure to a given pollutant in various microenvironments (e.g., indoor, 

outdoor, and in-vehicle microenvironments). One of the key strengths of the APEX model is its 

ability to estimate hourly exposures and doses for all simulated individuals in a sampled 

population. However, APEX is limited in that uncertainties in the predicted distributions (e.g., 

age, activity data, commuting patterns, personal activities) have not been addressed. The APEX 

model has not been evaluated, however, the pCNEM, a Canadian conceptional version of the 

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NEM model, has been evaluated for estimation of PM10. Since pCNEM is similar to the 

pNEM/NEM, the APEX model should work as well as pCNEM. (See discussion on pCNEM 

later is this section.) 

MENTOR/SHEDS is capable of simulating individuals exposures in various 

microenvironments (outdoors, residence, office, school, store, restaurant, bar, and vehicles) 

using spatial concentration data for each census tract. The indoor and in-vehicle pollutant 

concentrations are calculated using specific equations for the microenvironment and ambient 

pollutant concentration relationship. Randomly selected characteristics for a fixed number of 

individual are selected to match demographics within the census tract for age, gender, 

employment status, and housing type. Smoking prevalence statistics by gender and age is 

randomly selected for each individual in the simulation. Diaries for activity patterns are matched 

for the simulated individual by demographic characteristics (Burke et al., 2001). 

Zidek et al. (2000, 2003, 2005) described a methodology for predicting human exposure to 

environmental pollutants. The methodology builds on earlier models such as SHEDS and 

pNEM/NEM and provides a WWW platform for developing a wide variety of models. 

pCNEM, a platform model developed from this methodology, is a Canadian PC version of NEM. 

pCNEM was used to estimate a conditional predictive exposure distribution for PM10 in London. 

An important feature of pCNEM is it’s ability to estimate the effects of reductions in ambient 

levels of pollutants. 

An important source of uncertainty in existing exposure modeling involves the creation of 

multiday, seasonal, or year-long exposure activity sequences based on 1- to 3-day activity data 

for any given individual from CHAD. Activity pattern data sets in CHAD vary for different 

studies and may not be representative of the population. The commuting data are for home-towork and may not be representative of other commuting patterns. Correlations among human 

activities that can impact microenvironmental concentrations are not captured. See Annex AX3. 

Currently, appropriate longitudinal data are not available and the existing models use various 

rules to derive longer-term activity sequences utilizing 24-h activity data from CHAD. 

Of the above models, only NEM/pNEM have been used in the prior NAAQS review. The 

pNEM probabilistic model builds on the earlier NEM deterministic exposure model. The model 

takes into consideration the temporal and spatial distribution of people and O3 in the area of 

consideration, variations in O3 concentrations in the microenvironment, and the effects of 

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exercise-increased ventilation on O3 uptake. The pNEM models have been applied to nine urban 

areas and a summer camp. The models used activity data from the Cincinnati Activity Diary 

Study (CADS) along with time-activity data from several other studies. Data from stationary 

monitoring sites were used to estimate outdoor O3 exposure. Indoor O3 decay was assumed to be 

proportional to the indoor O3 concentration. An algorithm assigned the EVR associated with 

each exposure event. The EVR for the outdoor children model was generated using a module 

based on heart rate data by Spier et al. (1992) and Linn et al. (1992). 

Characterization of Exposure 

The use of ambient air monitoring stations is the most common surrogate for assigning 

exposure in epidemiological studies. Since the primary source of O3 exposure is the ambient air, 

monitoring concentration data would provide the exposure outdoors while exercising, a potential 

important exposure to evaluate in epidemiological studies. Monitored concentrations are useful 

for a relative assignment of exposure with time if the concentration were uniform across the 

region; the time-activity pattern were the same across the population; and the housing 

characteristics, such as ventilation rates and the O3 sinks contributing to its indoor decay rates, 

were constant for the study area. Since these factors vary by population and location there will 

be errors in the magnitude of the total exposure and in the relative total exposure assignment 

based solely on ambient monitoring data. 

Personal O3 exposure measurements have been made for potentially susceptible 

populations (children, outdoor workers, the elderly, and individuals with chronic obstructive 

pulmonary disease). Children and outdoor workers have somewhat higher exposures than other 

individuals because they spend more time outdoors engaged in moderate and heavy exertion. 

Children are also more active outside and, therefore, have a higher minute ventilation rate than 

most adults (Klepeis et al., 1996, 2001). Available exposure studies suggest trends in exposure 

magnitude for some populations, however, additional exposure studies are needed to generalize 

differences in exposure between the general population and potentially susceptible populations. 

Table 3-3 summaries the findings of available exposure studies. 

Ozone concentrations in various microenvironments under a variety of environmental 

conditions have been reported in the literature. In the absence of an indoor O3 source, 

concentrations of O3 indoors are lower than that found in the ambient air. Ozone concentrations 

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Table 3-3. Personal Exposure Concentrations 

Personal Exposure Meana 

Location, Population, Sample Duration n (range) (ppb) Reference 

San Diego, CA, Asthmatics ages 9-18 years, 12 12 ± 12 (0-84) Delfino et al. (1996) 

12 hour 10 weekend 

12 weekday 

Vancouver, Canada, Adult Workers, Daily 585 Brauer and Brook 

High indoor time (ND-9) (1997) 

Moderate indoor time (ND-12) 

Only outdoor (2-44) 

Southern California, Subjects 10-38 years 24 Liu et al. (1997) 

Spring 13.6 ± 2.5 (- to 80) 

Fall 10.5 ± 2.5 (- to 50) 

Montpellier, France, Adults, Hourly 16 34.3 ± 17.6 (6.5-88) Bernard et al. (1999) 

Winter 15.4 ± 7.7 (6.5-40) 

Summer 44.1 ± 18.2(11-88) 

Southern California, Children 6-12 years, 169 Geyh et al. (2000) 

$6 days 6.2 ± 4.7 (0.5-41) 

Upland - winter 19 ± 18 (0.5-63) 

- summer 5.7 ± 4.2 (0.5-31) 

Mountain - winter 25 ± 24 (0.5-72) 

- summer 

Baltimore, MD, Technician, Hourlyb 1 Chang et al. (2000) 

Winter 3.5 ± 7.5 (ND-49) 

Summer 15 ± 18 (ND-76) 

Baltimore, MD, Adults 75 ± 7 years, Daily 20 Sarnat et al. (2000) 

Winter 3.5 ± 3.0 (ND-9.9) 

Summer 0. ± 1.8 (ND-2.8) 

a ND = not detected.

b Measurements made following scripted activities for 15 days.

in microenvironments were found to be primarily controlled by ambient O3 concentrations and 

the AER: they increase with increasing AER. To a lesser extent, O3 concentrations in 

microenvironments are influenced by the ambient temperature, time of day, indoor 

characteristics (e.g., presence of carpeting), and the presence of other pollutants in the 

microenvironment. Table 3-4 describes the findings of the available studies. 

Factors Affecting Ozone Concentrations 

Ozone and other photochemical oxidants are formed in the ambient air from the reaction of 

sunlight with vehicle emissions, gasoline fumes, solvent vapors, and power plant and industrial 

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Table 3-4. Indoor/Outdoor Ozone Concentrations in Various Microenvironments 

Location and Ventilation Indoor/Outdoor Conditions Concentrations Comments 

 New England States (9) 20 ppb/40 ppb Schools represented a variety of environmental conditions - varying ambi

. Fall concentrations, sources, geographic locations, population density, traffic 

building types. Average O3 concentrations were low in the morning and p

during the early afternoon. O3 concentrations averaged for all schools mo

Mexico City, School 

Windows/Doors Open (27) 0 to 247 ppb/ Study conducted over 4 d period during winter months. Two-minute averWindows/Doors Closed Cleaner Off (41) 

64 to 361 ppb measurements were taken both inside and outside of the school every 30 m

10 a.m. to 4 p.m. Estimated air exchange rates were 1.1, 2.1, and 2.5 h!1 

Windows/Doors Closed Cleaner medium, and high flow rates. Ozone concentrations decreased with increOn (47) relative humidity. 

Mexico City 

Homes 5 ppb/27ppb (7 d) Ozone monitoring occurred between September and July. Study included 

7 ppb/37 ppb (14 d) and 145 homes. Most of the homes were large and did not have air condi

Ninety-two percent of the homes had carpeting, 13% used air filters, and 

Schools 22 ppb/73 ppb humidifiers. Thirty-five percent opened windows frequently, 43% somet

22% never between 10 a.m. and 4 p.m. Ozone monitored at schools sites from 8 a.m. to 1 p.m. daily for 14 consecutive days. Homes monitored focontinuous 24-h periods for 7 and 14 consecutive days. 

Boston, MA, Homes (9) 

Winter - continuously 0 to 20.4 ppb/4.4 to Study examined the potential for O3 to react with VOCs to form acid aero

24.5 ppb Carbonyls were formed. No clear trend of O3 with AERs. The average A

0.9 h!1 during the winter and 2.6 h!1 during the summer. Four residences 

Summer - continuously 0 to 34.2 ppb/8.2 to and nine in summer with over 24 h samples collected. 

51.8 ppb 

Los Angeles, Homes (239) 13 ppb/37 ppb Four hundred and eighty-one samples collected inside and immediately o

home from February to December. Concentrations based on 24-h averag

concentrations indoors and outdoors. Low outdoor concentrations resulteindoor concentrations. However, high outdoor concentrations resulted in indoor concentrations. 

Burbank, CA 

Telephone Switching Station 0.2 to 1.0/1.0–21.1 Major source of O3 was transport from outdoors. From early spring to lat

ppb concentrations peaked during the early afternoon and approach zero at su

AER ranged from 1.0 to 1.9 h!1 . 

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Table 3-4 (cont’d). Indoor/Outdoor Ozone Concentrations in Various Microenvironme

Location and Ventilation Indoor/Outdoor Conditions Concentrations Comments Montpellier, France 15.5/32.0 ppb Ozone measurements made over 5-d periods in and outside of 21 homes d

Homes (110) summer and winter months. The winter I/O ratio was 0.31 compared to 0

summer months. 

Southern CA, Homes Ozone measurements were taken at 119 homes (57 in Upland and 62 in to

Upland 11.8/48.2 ppb located in the mountains) during April and May. Concentrations were baMountains 2.8/35.7 ppb average monthly outdoor and average weekly indoor concentrations. Ind

on the home location, number of bedrooms, and presence/absence of an air conditioner. 

Krakow, Poland, Museums Ozone continuously monitored at five museums and cultural centers. Mo

Cloth Hall 3.2/25.7-27.4 ppb conducted during the summer months for 21 to 46 h or 28 to 33 days at eaMatejko 8.5/20.0 ppb Indoor concentration found to be dependent on ventilation rate, i.e., when Wawel Castle 2.5/14.7 ppb ventilation rate was high, the indoor O3 concentrations approached that of National 1.5/11.0 ppb O3. Rooms sequestered from outdoor air, or where air was predominantly 

through charcoal filters, O3 levels indoors greatly reduced. 

Buildings, Greece There was no heating/air conditioning system in the building at Thessalon

Thessalonki 9.39/15.48 ppb Windows were kept closed during the entire monitoring period. CompletAthens 8.14/21.66 ppb exchange took place every 3 h. The air conditioning system in continuou

the Athens site recirculated the air. Complete air exchange estimated to bMonitoring done for 30 days at each site, but only 7 most representative d

Patrol cars, NC 11.7/28.3 ppb Patrol cars were monitored Mon. through Thurs. between the hours of 3 p

midnight on 25 occasions during the months of Aug., Sept., and Oct. Ou

concentrations were taken from ambient monitoring station. Air inside thcar was recirculated cool air. 

University of CA <20 to 40 ppb/— Room volume was 40 m3 . Ozone concentrations increased proportionatel

Photocopy room increasing use of photocopier. 

Home/office BLD* to 290 ppb/— Room volume was 27 m3 . Doors and windows closed. Heating/air condi

O3 generators and mechanical ventilation systems off. Ozone generators operated for 9

High O3 concentrations noted when O3 generator used at high setting. AE

0.3 h!1 . Ozone concentrations varied depending on unit tested. 

3-66

*BLD = Below limit of detection 
emissions (See Chapter 2 for a discussion of O3 atmospheric chemistry). Ozone enters the 

indoor environment primarily through infiltration from outdoors through building components, 

such as windows, doors, and ventilation systems. There are also a few indoor sources of O3 

(photocopiers, facsimile machines, laser printers, and electrostatic air cleaners and precipitators) 

(Weschler, 2000). Generally O3 emissions from office equipment and air cleaners are low 

except under improper maintenance conditions. Reported O3 emissions from office equipment 

range from 1300 to 7900 μg/h (Leovic et al., 1996, 1998). Most air cleaners (particulate 

ionizers) emitted no or only a small amount (56 to 2757 μg/h) of O3 during operation (Niu et al., 

2001). Emissions from O3 generators can range from tens to thousands of micrograms per hour 

(Weschler, 2000; U.S. Environmental Protection Agency, 1996). 

Other photochemical oxidants (peroxyacyl nitrates; PAN and PPN) have no known direct 

emission sources indoors. Although not a significant source of indoor PAN. PAN can form in 

the indoor environment from the reaction of the OHA or NO3 with acetaldehyde to form the acetyl 

radical, CH3CO (Grosjean et al., 1996). The acetyl radical then reacts with oxygen to for an 

acetylperoxy radical which reacts with NO2 to form PAN. Peroxyacyl nitrates primarily occur in 

the indoor environment from infiltration through the building envelop and through openings in 

the building envelopment. 

The concentration of O3 in indoor environments is dependent on the outdoor O3 

concentration, the AER or outdoor infiltration, indoor circulation rate, and O3 removal processes 

through contact with indoor surfaces and reactions with other indoor pollutants. Since O3 

concentrations are generally higher during the warmer months, indoor concentrations will likely 

be highest during that time period. (See earlier discussion on ambient concentrations of O3.) 

Air exchange rates vary depending on temperature differences, wind effects, geographical 

region, type of heating/mechanical ventilation system, and building type (Weschler and Shields, 

2000; Colome et al., 1994). The balance of the flow of air in and out of a microenvironment is 

greatest in a residential building when a window or door is open (Johnson et al., 2004; HowardReed et al., 2002). The opening of windows or doors is dependent on the building occupancy, 

season, housing density, the presence of air conditioning, and wind speed (Johnson and Long, 

2004). When windows and doors are closed, the dominant mechanism controlling AERs is 

infiltration through unintentional openings in the building envelope. Williams et al. (2003a, 

2003b) reported AERs of 0.001 to 4.87 h!1

 in 37 homes in Research Triangle Park, NC. Chan 

3-67
et al. (2005) compared air leakage measurements for 70,000 houses. Older and smaller houses 

had higher normalized leakage areas than newer and larger houses. Meng et al. (2004) also 

attributed higher AERs to the age of the housing stock. AERs for homes in Houston, TX and 

Elizabeth, NJ were averaged for all four seasons, the highest AER, 1.22 h!1

, was noted for homes 

in Elizabeth, NJ where the homes were older. Evaluations of AERs for residential structures was 

reported by Murray and Burmaster (1995) and includes AERs for 2,844 residential structures in 

four different climatic regions by season (winter, spring, summer, and fall). The AER for all 

seasons across all regions was 0.76 h!1

 (arithmetic mean) (Region 1: IN, MN, MT, NH, NY1, 

VT, WI; Region 2: CO, CT, IL, NJ, NY2, OH, PA, WA; Region 3: CA3, MD, OR, WA; 

Region 4: AZ, CA4, FL, TX). The AERs were generally higher during the warm seasons, when 

ambient O3 concentrations are highest. Data for the warmest region during the summer months 

may not be representative of all homes because measurements were made in southern California 

where windows are open and air conditioning is not used. 

Average mean (median) AERs of 2.45 (2.24), 1.35 (1.09), and 2.22 (1.79) h!1

 were 

reported by Lagus Applied Technology, Inc. (1995) for schools, offices, and retail 

establishments in California. Mean AERs for schools, offices, and retail establishments in 

Oregon and Washington were 0.32, 0.31, and 1.12 h!1

 (Turk et al., 1989)—considerably less than 

that reported by Lagus Applied Technology. Park et al. (1998) reported mean AERs ranging 

from 1.0 to 47.5 h!1

 for stationary vehicles under varying ventilating conditions. Where 

available, AERs for other studies are included in Table 3-10. 

The most important removal process for O3 in the indoor environment is deposition on, and 

reaction with, indoor surfaces. The rate of deposition is material-specific. The removal rate will 

depend on the indoor dimensions, surface coverings, and furnishings. Smaller rooms generally 

have larger surface-to-volume ratio (A/V) and remove O3 faster than larger rooms. Fleecy 

materials, such as carpets, have larger surface-to-volume ratios and remove O3 faster than 

smooth surfaces (Weschler, 2000). However, the rate of O3 reaction with carpet diminishes with 

cumulative O3 exposure (Morrison and Nazaroff, 2000, 2002). Weschler (2000) compiled the O3 

removal rates for a variety of microenvironments. Generally, the removal rates ranged between 

3.0 and 4.3 kd (A/V)/h–1. The highest removal rate, 7.6 kd (A/V)/h–1, was noted for a clean room 

(Weschler et al, 1989). 

3-68
Ozone chemical reactions in the indoor environment are analogous to those reactions 

occurring in the ambient air (See discussion on atmospheric chemistry in Chapter 2). Ozone 

reacts with unsaturated VOCs in the indoor environment, primarily terpenes or terpene-related 

compounds from cleaning products, air fresheners, and wood products. The reactions are 

dependent on the O3 indoor concentration, the indoor temperature and, in most cases, the air 

exchange rate/ventilation rate. It is beyond the scope of this document to discuss actual 

concentrations of O3 reaction products in the indoor environment; however, many of the reaction 

products may more negatively impact human health and artifacts in the indoor environment than 

their precursors (Wolkoff et al., 1999; Wilkins et al., 2001; Weschler et al., 1992; Weschler and 

Shields, 1997; Rohr et al., 2002; Nøjgaard et al., 2005). Primary reaction products are Criegee 

biradicals, nitrate radicals, and peroxyacetyl radicals. Secondary reaction products are hydroxy, 

alkyl, alkylperoxy, hydroperoxy, and alkoxy radicals. Reactions with alkenes can produce 

aldehydes, ketones, and organic acids (Weschler and Shields, 2000; Weschler et al., 1992). 

Hydroxyl radicals formed from the reaction of O3 with VOCs, nitric oxide and 

hydroperoxy, and other intermediate products can react with nitrogen compounds, sulfur 

dioxide, and carbon monoxide to produce significantly more toxic compounds (Sarwar et al., 

2002; Orzechowska and Paulson, 2002; Fick et al., 2003, 2004; Van den Bergh et al., 2000; Fan 

et al., 2003; Wilkins et al., 2001; Clausen et al., 2001; Rohr et al., 2002, 2003; Poupard et al., 

2005; Blondeau et al., 2005). The reaction between O3 and terpenes also has been shown to 

increase the concentration of indoor particles (Weschler and Shields, 1999, 2003; Weschler, 

2004; Clausen et al., 2001; Fan et al, 2003; Wainman et al., 2000), possibly from further 

reactions of the hydroxy radical with terpenes (Sarwar et al., 2002). 

Decomposition and formation of PAN in the indoor environment are influenced by NO2 

and NO. Decomposition of PAN is expected to be a relatively fast process when indoor O3 

levels are low and when motor vehicle emissions are large or there is an indoor source of NOx 

(Weschler and Shields, 1997). 

Factors Affecting the Relationship between Ambient Concentrations and Personal 

Exposures to O3 

Ambient O3 concentrations vary with the time of day, season of the year, and among 

locations. Personal exposure to O3 is influenced by the microenvironmental concentration and 

the amount of time spent in each microenvironment. The majority of the population spends, on 

3-69
average, nearly 90% of their time in an indoor microenvironment. Since there are few indoor 

sources of O3, O3 ambient concentration may be the most important factor that affects average 

population exposure in the indoor environment. 

Indoor O3 concentrations also are affected by several other factors and mechanisms. 

Studies have shown that in addition to the ambient O3 concentrations, indoor O3 concentrations 

are influenced by the air exchange rate or outdoor infiltration, increasing with increasing air 

exchange. Once indoors, the O3 concentration is affected by the indoor circulation rate and O3 

removal through contact with indoor surfaces and reactions with other indoor pollutants. 

Studies on the effect of elevation on O3 concentrations found that concentrations increased 

with increasing elevation (Väkevä et al., 1999; Johnson, 1997). Since O3 monitors are frequently 

located on rooftops in urban settings, the concentrations measured there may overestimate the 

exposure to individuals outdoors in streets and parks, locations where people exercise and their 

maximum O3 exposure is more likely to occur. 

In epidemiologic studies investigating acute and chronic health outcomes using ambient 

monitoring data from stationary monitoring sites, O3 exposure assessment was affected by the 

distance between home and the monitoring site, gender, time-activity patterns (e.g., percentage 

of time spent outdoors, type of outdoor activity, time of day during outdoor activity), and indoor 

air exchange rates (e.g., ventilation conditions, home characteristics) (Geyh et al., 2000; Lee 

et al., 2002, 2004; Liu et al., 1995, 1997; Chang et al., 2000; Chan et al., 2005; O’Neill et al., 

2003; Brauer and Brook, 1997). Geyh et al. (2000) observed higher indoor and personal O3 

concentrations in a southern California community with 2% air-conditioned homes compared to 

a community with 93% air-conditioned homes during the summer (high O3) months, but showed 

no difference in O3 levels during the winter (low O3) months. People that work outdoors tend to 

be exposed to higher levels of O3 (Brauer and Brook, 1997; O’Neill et al., 2003). Lee et al. 

(2004) observed that personal O3 exposure was positively correlated with outdoor time (r = 0.19, 

p < 0.01) and negatively correlated with indoor time (r = !0.17, p < 0.01). Additional factors 

that affected indoor O3 levels were air conditioning, window fans, and window opening. The O3 

exposure assessment study by Liu et al. (1995) found that after adjusting for time spent in 

various indoor and outdoor microenvironments (e.g., car with windows open, car with windows 

closed, school, work, home, outdoors near home, outdoors other than near home), mean 12-hour 

ambient O3 concentrations explained 32% of the variance in personal exposure in the summer. 

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In a southern California study by Avol et al. (1998a), boys were found to spend more time 

outdoors and be more physically active than girls. Another southern California study found that 

boys were outdoors 30 minutes longer than girls and had higher personal O3 exposure during 

both high and low O3 months (Geyh et al., 2000). 

The announcement of smog alerts or air quality indices may influence personal exposures 

to O3 by causing individuals to alter behaviors (avoidance behavior). Neidell (2004), in his 

evaluation of the effect of pollution on childhood asthma, examined the relationship between the 

issuance of smog alerts or air quality indices for several counties in California and hospital 

admissions for asthma in children under age 18 years (not including newborns). Smog alerts are 

issued in California on days when O3 concentrations exceed 200 ppb. There was a significant 

reduction in the number of asthma-related hospital admissions in children ages 1 to 12 years on 

smog alert days, indicating that avoidance behavior might be present on days of high O3 

concentrations. Changes in population behavior as a function of concentration complicate the 

estimation of health effects from population-based studies; thus, it may be desirable to include 

sensitivity analyses that eliminate high O3 days, particularly in areas where avoidance behavior 

is expected. 

Potential Sources of Error Resulting from the Use of Ambient Ozone Concentrations in 

Epidemiological Analyses 

There is no clear consensus among exposure analysts as to how well stationary monitor 

measurements of ambient O3 concentrations represent a surrogate for personal O3 exposure. 

The microenvironmental (indirect) approach and the personal sampling (direct) approach 

(Navidi et al., 1999; Ott, 1982, 1985) have been used to assess personal exposure in air pollution 

epidemiologic studies; however, both approaches are associated with measurement error. 

To determine personal exposure using the microenvironmental approach, the concentrations of 

the various microenvironments are multiplied by the time spent in each microenvironment. 

Both the concentration and time component contribute to the measurement error. There is no 

time component to the measurement error in the personal sampling approach, however, the 

estimation of exposure using personal monitoring devices contributes to measurement error, 

especially in the case of O3. Passive badges are commonly used for monitoring O3 integrated 

personal exposure. Their sensitivity to wind velocity, badge placement, and interference with 

other copollutants may result in measurement error. 

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Results from the error analysis models developed by Navidi et al. (1999) using two 

different statistical designs (bidirectional case-crossover and multi-level analytic designs) 

indicated that neither the microenvironmental nor personal sampling approach gave reliable 

health effect estimates when measurement errors were uncorrected. The nondifferential 

measurement error biased the effect estimates toward zero under the model assumptions. 

However, if the measurement error was correlated with the health response, a bias away from the 

null could result. The use of central ambient monitors to estimate personal exposure has a 

greater potential to introduce bias since most people spend the majority of their time indoors, 

where O3 levels tend to be much lower than outdoor ambient levels. If the error is of a fixed 

amount (i.e., absolute differences do not change with increasing concentrations), there is no bias. 

However, if the error is not a fixed difference, this error will likely attenuate the O3 risk estimate 

as personal O3 exposures are generally lower than ambient O3 concentrations. 

Several studies have examined relationships between measured ambient O3 concentrations 

from fixed monitoring sites and personal O3 exposure (Avol et al., 1998b; Brauer and Brook, 

1995, 1997; Chang et al., 2000; Delfino et al., 1996; Lee et al., 2004; Liard et al., 1999; Linn 

et al., 1996; Liu et al., 1995, 1997; O’Neill et al., 2003; Sarnat et al., 2001). Two studies by 

Sarnat et al. (2001, 2005) examined relationships between individual variations in personal 

exposure and ambient O3 concentrations. In the first study conducted in Baltimore, MD, the 

association between 24-h average ambient O3 concentrations from a centrally-located monitoring 

site and 24-h average personal O3 exposures was evaluated in a cohort of older adults (n = 20), 

individuals with COPD (n = 15), and children (n = 21) (Sarnat et al., 2001). Personal exposures 

were measured repeatedly for each subject for a total of 800 person-days; thus, analyses of 

personal exposure data were conducted using mixed models with subjects modeled as random 

variables to account for between-subject variation. The mixed regression effect estimates 

were $ = 0.01 (95% CI: !0.01, 0.03) for the summer (196 paired samples) and $ = 0.00 for the 

winter (449 paired samples). However, in the second study conducted in Boston, MA with a 

cohort of 20 healthy senior citizens and 23 school children, significant associations were 

observed between 24-h average ambient O3 concentrations and 24-h average personal O3 

exposures (Sarnat et al., 2005). The mixed regression effect estimates were $ = 0.27 (95% CI: 

0.18, 0.37) and $ = 0.04 (95% CI: 0.00, 0.07) during the summer (332 paired samples) and 

winter (288 paired samples), respectively. In the Boston study, the regression coefficients 

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indicated that ambient O3 concentrations were predictive of personal O3 exposures; however, 

ambient O3 levels overestimated personal exposures 3- to 4-fold in the summer and 25-fold in 

the winter. 

Important predictors for personal O3 exposure were examined (Xue et al., 2005) in the 

Harvard Southern California Chronic Ozone Exposure Study. In addition to ambient O3 

concentrations from central monitors, indoor and outdoor O3 concentrations at the participant’s 

homes, and personal O3 concentrations were measured using passive O3 samplers. During a 

one-year period, 160 children from two southern California communities (Upland and the 

San Bernardino mountains) were monitored for six consecutive days for at least six of the 

12 months. Outdoor O3 concentrations at the participant’s homes were very similar to 

central ambient O3 concentrations, and personal O3 concentrations were close to indoor O3 

concentrations. At different time points throughout the year, the average ratios of personal to 

central O3 concentrations were relatively stable, being around 0.3 (SD 0.13). Central O3 

concentrations were approximately 3 times higher compared to personal concentrations during 

the O3 season (May to September) and approximately 5 times higher during the non-O3 season 

(October to April). Xue et al. (2005) found that ambient O3 concentrations from central 

monitors, after adjusting for time-activity patterns and housing characteristics from 

questionnaire data, reasonably predicted personal O3 concentrations; a 1 ppb increase in 

ambient O3 concentration was associated with a 0.54 ppb (95% CI not provided) increase in 

personal O3 exposure (Pearson r = 0.76). The regression coefficient for the relationship between 

ambient O3 concentrations and personal O3 exposures without adjustment for time-activity and 

housing characteristics, which is of most relevance to epidemiologic time-series studies, was 

not presented. 

Chang et al. (2000) compared 1-h personal and ambient O3 measurements in older adults in 

various microenvironments in Baltimore, MD, using activity data from the National Human 

Activity Pattern Survey study (Klepeis, 1999). Activities were scripted to simulate activities 

performed by older adults (65+ years of age). In total, 180 1-h personal samples were collected 

in each season (summer and winter). There was no correlation between personal and ambient O3 

concentrations in the indoor residence (r = 0.09 in summer and r = 0.05 in winter), although a 

moderate correlation was found in other indoor environments such as restaurants, hospitals, and 

shopping malls (r = 0.34 and r = 0.46 for summer and winter, respectively). In comparison, the 

3-73
correlation in outdoor environments (near and away from roads) was moderate to high (0.68 #

r # 0.91) and statistically significant. Regression coefficients for the relationship between 

personal and ambient O3 concentrations were not reported in this study. 

Brauer and Brook (1995, 1997) observed that the daily averaged personal O3 measurements 

and ambient concentrations were well-correlated after stratifying groups by time spent outdoors. 

Clinic workers (n = 25; 24-hour samples), teenage camp counselors (n = 25; 24-hour samples), 

and farm workers (n = 15; 6-14 h work shift samples) spent 0 to 25%, 7.5 to 45%, and 100% of 

their monitored time outdoors, respectively. The personal to ambient O3 concentration ratios 

were significantly different for the clinic workers (0.28) and farm workers (0.96). Ambient O3 

concentrations and time spent outdoors explained more of the variability in the personal O3 

measurements for outdoor farm workers compared to the clinical workers. However, the 

Spearman correlation coefficients were comparable, 0.60 and 0.64 for the clinic workers and 

farm workers, respectively, indicating that even though the clinic workers spent considerable 

amounts of time indoors or in transit there was still reasonable correlation between the day-today variations in mean personal O3 exposures and mean O3 concentrations. 

A study by O’Neill et al. (2003) examined 107 pairs of ambient and personal O3 

measurements from 39 outdoor workers in Mexico City using a longitudinal analysis method. 

Two to seven personal measurements were collected on each of the 26 monitoring days, which 

were averaged and then compared with the ambient concentrations. They estimated that a 1 ppb 

increase in ambient O3 concentration was associated with a 0.56 ppb (95% CI: 0.43, 0.69) 

increase in personal O3 concentration. In a Paris, France study by Liard et al. (1999), adults 

(n = 55) and children (n = 39) wore passive O3 monitors for 4 consecutive days during three 

periods. For each period, all adults wore the O3 monitors over the same 4 days. Likewise, all 

children wore monitors over the same 4 days for each of the three periods, but on different days 

from the adults. The ambient O3 concentrations from the stationary monitoring sites did not 

explain a high percentage of the variance of personal O3 exposure (nonsignificant [value not 

stated] in adults and 21% in children). However, when personal measurements from all subjects 

were aggregated for each of the six periods, the 4-day mean personal O3 exposure was found to 

be highly correlated with the corresponding mean ambient concentration (r = 0.83, p < 0.05). 

Similarly, a study of Los Angeles school children by Linn et al. (1996) found that daily 24-h 

3-74
average ambient O3 concentrations from a central site were well-correlated (r = 0.61) with daily 

averaged personal O3 exposures. 

The relationship between personal O3 exposures and ambient O3 concentrations from 

central monitoring sites differed by city in the two studies by Sarnat et al. (2001, 2005). In the 

Baltimore study (Sarnat et al., 2001) no association was observed, while in the Boston study 

(Sarnat et al., 2005) ambient O3 concentrations were significantly associated with personal O3 

exposures in mixed effects regression models. These results suggest that O3 concentrations 

measured at central ambient monitors may explain, at least partially, the variance of individual 

personal exposures; however, this relationship is influenced by factors such as air exchange rates 

in housing and time spent outdoors, which may vary by city. This was further supported by 

results from the southern California study by Xue et al. (2005). They found that incorporating 

information on time-activity and housing characteristics allowed reasonable prediction of 

personal O3 exposures from ambient O3 concentrations. 

Other studies observed that the daily averaged personal O3 exposures from the population 

were well correlated with monitored ambient O3 concentrations, although substantial variability 

existed among the personal measurements. In other words, centrally-located ambient O3 

monitors are likely to be representative of day-to-day changes in O3 exposure experienced by the 

population. Brauer and Brook (1997) indicated that these results have implications for large 

epidemiologic studies which often depend upon fixed-site outdoor ambient monitors to estimate 

exposures. It should be noted, however, that although there are correlations between aggregate 

personal and monitored ambient O3 concentrations, the absolute personal concentrations are 

generally considerably lower than the monitored ambient O3 concentrations. 

In summary, results indicate that the relationship between ambient O3 concentrations and 

personal exposure will vary depending on individual- or city-specific factors such as timeactivity patterns, indoor air exchange rates, and housing conditions, creating potential 

measurement errors. The expectations based on statistical modeling considerations are that these 

exposure measurement errors or uncertainties will decrease estimates of effect in O3 health 

effects analyses, making it difficult to detect a true underlying association between the exposure 

and the health outcome studied. Also of special concern in interpreting results from mortality 

and hospitalization time-series studies is to what extent the ambient O3 concentrations are 

representative of personal O3 exposures in a particularly susceptible group of individuals, the 

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debilitated elderly, as the correlation between the two measurements have not been examined in 

this population. However, until more data on O3 exposure become available, the use of 

monitored ambient O3 concentrations as a surrogate for exposures is not expected to change the 

principal conclusions from O3 epidemiologic studies. 

Exposure to Related Photochemical Oxidants 

A variety of related photochemical oxidants produced outdoors, such as PAN and 

peroxypropionyl nitrate (PPN), can infiltrate into indoor environments. These compounds are 

thermally unstable and decompose to peroxacyl radicals and NO2. Exposure to related 

photochemical oxidants has not been measured, nor are these compounds routinely monitored at 

stationary monitoring sites. Available monitored concentrations of related photochemical 

oxidants may be found in Annex AX3. 

3.9 SUMMARY OF KEY POINTS 

The median of the daily maximum 8-h O3 concentration averaged over May to September 

2000 to 2004 was about 0.049 ppm. The daily maximum 1-h O3 concentrations could have been 

much higher in large urban areas or in areas downwind of large urban areas. For example, in 

Houston, TX, the daily maximum 1-h O3 concentrations have approached 0.20 ppm during this 

period. 

Daily maximum 8-h average O3 concentrations are lower than the maximum 1-h O3 

concentrations, but they are highly correlated. Within individual MSAs, O3 concentrations tend 

to be well correlated across monitoring sites. However, there can be substantial variations 

in O3 concentrations. Ozone in city centers tends to be lower than in regions either upwind or 

downwind because of titration by NO emitted by motor vehicles. 

Ozone concentrations tend to peak in early- to mid-afternoon in areas where there is strong 

photochemical activity and later in the day in areas where transport is more important in 

determining the O3 abundance. Summertime maxima in O3 concentrations occur in areas in the 

United States where there is substantial photochemical activity involving O3 precursors emitted 

from human activities. Monthly maxima can occur anytime from June through August. 

However, springtime maxima are observed in national parks, mainly in the western United States 

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and at a number of other relatively unpolluted monitoring sites throughout the Northern 

Hemisphere. For example, the highest O3 concentrations at Yellowstone NP tend to occur 

during April and May. Generally, monthly minima O3 concentrations tend to occur from 

November through February at polluted sites and during the fall at relatively remote sites. 

Nationwide, daily maximum 8-h O3 concentrations have decreased at the upper end of the 

distribution from 1990 to 2004. However, the daily maximum 8-h O3 concentrations toward the 

center of the distribution have not reflected these changes. Trends have not been consistent at 

national park sites; with downward trends observed at some sites and upward or no trends 

observed at others. At some sites, trends reversed direction in going from the 98th to the 95th 

percentile values. 

Sufficient data are not available for other atmospheric oxidants (e.g., H2O2, PAN) and 

oxidation products (e.g., HNO3, H2SO4) to relate concentrations of O3 to these species for use in 

time series studies. Data for these species are only obtained as part of specialized field studies. 

In general, secondary species, such as HNO3, H2SO4, H2O2, and PAN, are expected to be at least 

moderately correlated with O3. On the other hand, primary species are expected to be more 

highly correlated with each other than with secondary species, provided that the primary species 

originate from common sources. Concentrations of other oxidants are much lower than for O3 

and range from #1% for oxidants in particles to several percent for gas phase species. The 

relationship of O3 to PM2.5 is complex, because PM is not a distinct chemical species but is a mix 

of primary and secondary species. PM2.5 concentrations were positively correlated with O3 

during summer, but negatively correlated with O3 during winter at Ft. Meade, MD. PM10 

concentrations show similar relations with O3. 

PM

Co-occurrences of O3 (defined when both pollutants are present at an hourly average 

concentration of $0.05 ppm) with NO2 and SO2 are rare. For example, there were fewer than 

10 co-occurrences with either NO2 or SO2 in 2001. The number of co-occurrences for O3 and 

2.5 (defined as an hourly average O3 concentration $0.05 ppm and a 24-h average PM2.5 

concentration $40 μg/m3 occurring during the same 24-h period) also tended to be infrequent 

(<10 times) at most sites, but there were up to 20 such co-occurrences at a few sites. 

Policy relevant background O3 concentrations are used for assessing risks to human health 

associated with O3 produced from anthropogenic sources in continental North America. Because 

of the nature of the definition of PRB concentrations, they cannot be directly derived from 

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monitored concentrations, instead they must be derived from modeled estimates. Current model 

estimates indicate that ambient air PRB concentrations in the United States are generally 

0.015 ppm to 0.035 ppm. They decline from spring to summer and are generally <0.025 ppm 

under conditions conducive to high O3 episodes. However, PRB concentrations can be higher, 

especially at elevated sites during spring, due to enhanced contributions from hemispheric 

pollution and stratospheric exchange. 

Ozone exposure changes as a function of time of day, season, and microenvironment. 

Ambient O3 concentrations are generally higher during warmer seasons and during the weekday, 

peaking during the later portion of the day. Ozone concentrations in indoor microenvironments 

are generally lower than those concentrations encountered in the ambient air. There are few 

indoor sources of O3. Ozone occurs in indoor microenvironments primarily through infiltration 

through the building envelop and through windows, doors, and ventilation systems. The indoor 

O3 concentration is dependent on the outdoor concentration, the AER, indoor circulation rate, 

and removal processes. Consequently, measured and modeled exposures should take into 

consideration O3 diurnal weekly and seasonal variability and varying microenvironmental 

concentrations. 

Once indoors, O3 reacts with indoor surfaces, including surface coverings and furnishings. 

Ozone also will react with VOCs in indoor environments, primarily terpenes or terpene-related 

compounds. Ozone reactions with pollutants indoors are analogous to those reactions occurring 

in the ambient air, potentially exposing subjects to compounds significantly more toxic than O3. 

The available approaches for measuring personal O3 exposure include the direct approach, 

using a PEM, and the indirect approach, which measures or models exposure in the 

microenvironments the individual encounters. Both approaches are associated with 

measurement errors. 

There are difficulties in identifying chemical trapping agents for PEMs that can react 

with O3, and PEMs are sensitive to wind velocity, badge placement, and interference with other 

copollutants. Studies using PEMs show personal O3 exposures below those concentrations 

measured at stationary monitoring sites, when measurements are not adjusted for time spent 

outdoors, housing characteristics, age, gender, and occupation. 

The use of measured O3 concentrations from stationary ambient monitoring sites as 

surrogates for personal exposure may be affected by the O3 ambient concentration, percentage of 

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time spent outdoors, and type of outdoor activity. Epidemiologic studies investigating health 

outcomes using data from stationary monitoring sites found O3 exposure to be affected by the 

distance between the subjects’ location and the stationary monitor, individual activity patterns, 

and the O3 concentration in the microenvironment. 

The use of exposure models to evaluate O3 exposure to large populations over time is 

complicated by the fact that O3 is a secondary pollutant with complex nonlinear and multiscale 

dynamics in space and time. The existing comprehensive inhalation exposure models (NEM, 

pNEM, MENTOR/SHEDS, APEX, pCNEM) treat human activity patterns as sequences of 

exposure events. Estimates of activity levels are assigned from CHAD, the Consolidated Human 

Activities Database. 

Ambient O3 concentrations are estimated using emissions-based mechanistic models or 

ambient-data-based models. Models for estimating microenvironmental concentrations include 

the empirical, mass balance, and detailed CFD models. Mass balance modeling is the most 

common modeling approach to estimating concentrations in enclosed microenvironments. 

The pNEM/O3 population exposure model, the model used more extensively in O3 exposure 

modeling, includes a sophisticated mass balance model for indoor and vehicle 

microenvironments. There are three versions of the pNEM/O3 model: the general population, 

outdoor workers, and outdoor children. 

Results from O3 exposure studies indicate that the relationship between ambient O3 

concentrations and personal exposure/dose will vary depending on O3 concentrations and time 

spent in the various microenvironments, particularly the time spent outdoors where O3 

concentrations tend to be higher, and the personal activity level. Consequently, the O3 

exposure/dose may differ from the concentrations measured at stationary monitoring sites. 

However, until more data on O3 exposure become available, the use of monitored ambient O3 

concentrations as a surrogate for exposures is not expected to change the principal conclusions 

from O3 epidemiologic studies using community average health and pollution data. 

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4. DOSIMETRY, SPECIES HOMOLOGY, SENSITIVITY,

AND ANIMAL-TO-HUMAN EXTRAPOLATION

4.1 INTRODUCTION 

The dosimetry of ozone (O3) in humans has been examined in a series of studies published 

in the past decade. These studies further characterize the dose of O3 delivered to various sites in 

the respiratory tract (RT). Ozone, classified as a reactive gas, interacts with surfactant, 

antioxidants, and other compounds in the epithelial lining fluid (ELF). Researchers have 

attempted to obtain a greater understanding of how these complex interactions affect O3 uptake 

and O3-induced injury. New work has also been completed evaluating species differences in 

responses to O3 exposures, which allow more accurate quantitative extrapolation from animals 

to humans. 

This chapter is not intended to be a complete overview of O3 dosimetry and animal-tohuman comparisons, but rather, it is an update of the dosimetry/extrapolation chapter from the 

last O3 criteria document (U.S. Environmental Protection Agency, 1996), or 1996 O3 AQCD, and 

other reviews of the earlier published literature. The framework for presenting this chapter is 

first a discussion in Section 4.2 of general concepts of the dosimetry of O3 in the RT. Bolusresponse studies are then presented in Section 4.2.1 followed by general uptake studies in 

Section 4.2.2. Dosimetry modeling is presented in Section 4.2.3 followed by the summary and 

conclusions for the dosimetry material in Section 4.2.4. The chapter continues in Section 4.3 

with a discussion of species comparisons and ends with a discussion of animal-to-human 

extrapolation. More detailed discussions of the studies are presented in the supporting material 

to this chapter (Annex AX4). The toxicological effects of O3 in laboratory animals and in vitro 

test systems are discussed in Chapter 5 and direct effects of O3 in humans are discussed in 

Chapter 6. The historical O3 literature is very briefly summarized in this chapter, providing a 

very concise overview of previous work. The reader is referred to the 1996 O3 AQCD for more 

detailed discussion of the literature prior to the early 1990s. 

4-1
4.2 DOSIMETRY OF OZONE IN THE RESPIRATORY TRACT 

Ozone dosimetry refers to the measurement or estimation of the amount of O3 or its 

reaction products reaching and persisting at specific sites in the RT following an exposure. The 

compound most directly responsible for toxic effects may be the inhaled gas O3 or one of its 

chemical reaction products. Complete identification of the actual toxic agents and their 

integration into dosimetry is a complex issue that has not been resolved. Dosimetric studies 

attempt to quantify the amount of O3 retained in the lung (i.e., not exhaled) or the dose of O3 or 

its active metabolites (e.g., aldehydes or peroxides) delivered to target cells or tissues (i.e., dose 

per cell or tissue surface area). For comparison, epidemiologic studies may simply consider 

exposure concentration while clinical studies may consider the total amount of O3 inhaled 

(product of exposure concentration, duration, and minute ventilation). Hence, dosimetric studies 

seek to accurately quantify dose to target lung regions or tissues, whereas epidemiologic and 

clinical studies typically consider exposures. 

Figure 4-1 illustrates the structure of the lower airways with progression from the large 

airways to the alveolus. Understanding dosimetry as it relates to O3-induced injury is complex 

due to the fact that O3 interacts primarily with the ELF, which contains surfactant and 

antioxidants. Reactive products created by O3 can diffuse within the lung or be transported out 

of the lung to generate systemic effects. Antioxidant enzymes are the primary cellular defense 

against reactive species created by O3. The level and type of antioxidants varies between 

species, regions of the RT itself, and can be altered by O3 exposure. 

A considerable number of dosimetric studies were summarized in the 1996 O3 AQCD. 

These studies provided estimates of absorbed O3 in the RT as a whole or in regions such as the 

upper airways (URT) or lower airways (LRT), defined as being proximal or distal to the tracheal 

entrance, respectively. Estimates were obtained for both humans and animals via direct 

measurement and mathematical modeling. The mathematical models also estimated O3 doses to 

specific target sites such as the proximal alveolar region (PAR; first generation distal to the 

terminal bronchioles), which is also referred to as the centriacinar region (CAR; junction of 

conducting airways and gas exchange region) in some studies. 

4-2
Figure 4-1. Structure of lower airways with progression from the large airways to the 

alveolus. Panel (a) illustrates basic airway anatomy. Structures are epithelial 

cells, EP; basement membrane, BM; smooth muscle cells, SM; and 

fibrocartilaginous coat, FC. Panel (b) illustrates the relative amounts of liquid, 

tissue, and blood with distal progression. In the bronchi there is a thick liquid 

lining over a relatively thick layer of tissues. Even highly soluble materials 

moving from the air into the liquid layer have minimal systemic access via the 

blood. With distal progress, the protective liquid lining diminishes allowing 

increased access of compounds crossing the air-liquid interface to the tissues 

and the blood. 

Source: Panel (a) reproduced with permission (Weibel, E. R. [1980] Design and structure of the human lung. 

In: Fishman, A. P., ed. Pulmonary Diseases and Disorders. New York, NY: McGraw-Hill; p. 231). 

4-3
In general, the consensus of experimental and modeling studies summarized in the 

1996 O3 AQCD supported the following conclusions: (1) for the URT, animal and human 

studies suggested that O3 uptake is greater in the nose than the mouth but the effect of flow on 

uptake was equivocal; (2) for the LRT, predicted tissue doses (O3 flux to liquid-tissue interface) 

were very low in the trachea, increased to a maximum in the terminal bronchioles or first airway 

generation in the pulmonary region, and rapidly decreased with distal progression; (3) increasing 

tidal volume (VT) increases O3 uptake, whereas, increasing flow or breathing frequency (fB) 

decreases O3 uptake; (3) increasing flow shifts O3 uptake to the smaller peripheral airways, i.e., 

toward the CAR; and (4) similarly, the effect of exercise is to significantly increase the 

pulmonary region total dose (mass of O3) and the CAR dose (mass per unit surface area). 

Some cross-species in vivo comparisons were described in the 1996 O3 AQCD. 

For instance, comparing bronchoalveolar lavage (BAL) cells from rats and humans, it was 

estimated that a 0.4 ppm O3 exposure in exercising humans gave 4 to 5 times the O3 dose 

(retained) relative to rats exposed at rest to the same concentration. In vitro dosimetry studies in 

the 1996 O3 AQCD using isolated lung preparations showed that uptake efficiency is chemicalreaction dependent, indicating the importance of reaction product formation. These reaction 

products, created mainly by the ozonolysis of polyunsaturated fatty acids, included hydrogen 

peroxide, aldehydes, and hydroxyhydroperoxides, which are mediators of O3 toxicity. Other 

products are created by the reaction of O3 with other ELF constituents, all of which must be 

considered in understanding the dosimetry of O3. 

The next two sections (4.2.1 and 4.2.2) review the available new experimental studies 

on O3 dosimetry. Table AX4-1 in Annex AX4 summarizes the new human studies. 

4.2.1 Bolus-Response Studies in Humans 

The bolus-response method has been used by the Ultman group as an approach to explore 

the distribution of O3 absorption in the airways of humans. This noninvasive method consists 

of an injection of a known volume and concentration of O3 at a predetermined point during 

inspiration. Ozone uptake is the amount of O3 absorbed during a breath relative to the amount 

contained in the inhaled bolus. Figure 4-2 illustrates the uptake of a series of O3 boli as a 

4-4
V

Figure 4-2. Ozone uptake fraction as a function of volumetric penetration (VP) in a 

representative subject. Each point represents the O3 uptake of a bolus inspired 

by the subject. The volumes, VUA and VD, are the volume of the upper airways 

and anatomical dead space, respectively, and VP50% is the VP at which 50% of 

the inspired bolus was absorbed. In 47 healthy subjects (24 M, 23 F), Ultman 

et al. (2004) found that VP50% was well correlated with VD (r = 0.57, p < 0.001) 

and better correlated with the volume of the conducting airways, i.e., VD minus 

UA, (r = 0.65, p = 0.001). 

Source: Adapted from Ultman et al. (2004). 

function of volumetric penetration (VP), i.e., the volume between the center of mass of an 

inhaled bolus and the end of inspiration. The inspired O3 boli (for which the uptake fractions are 

illustrated in Figure 4-2) were 20 mL of 2 ppm O3. Kabel et al. (1994) have previously shown 

that varying the O3 concentration of inspired boli between 0.4 and 4 ppm does not affect the 

distribution of uptake as a function of VP. 

The O3 bolus-technique was used by Bush et al. (1996a) to ascertain differences in lung 

anatomy and gender that can alter the exposure-dose cascade. Forced vital capacity (FVC), total 

lung capacity (TLC) and anatomic dead space (VD) were determined for ten male and ten female 

subjects. Differences between subjects in absorption of a 20 mL O3 bolus injected into the 

inhaled breath could be explained by differences in VD. In particular, they concluded that the 

4-5
intrinsic mass transfer parameter (Ka) was proportional to the ratio of the respiratory flow to VD. 

In a subsequent study, Ultman et al. (2004) showed that the volume at which 50% of an 

inspired O3 bolus is absorbed was better associated with the volume of the lower conducting 

airways than VD (see Figure 4-2). Bush et al. (1996a) pointed out that the applicability of their 

results may be limited because of their assumptions that Ka was independent of location in the 

RT and that there was no mucous resistance. They further suggested that the dependence 

of Ka on flowrate and VD be restricted to flowrates #1000 mL/s until studies at higher rates 

have been performed. 

Nodelman and Ultman (1999) demonstrated that the uptake distributions of O3 boli were 

sensitive to the mode of breathing and to the airflow rate. As flowrates increased from 150 to 

1000 mL/s, O3 penetrated deeper into the lung and penetration was further increased by oral 

relative to nasal breathing. The authors suggest that the switch from nasal to oral breathing 

coupled with increases in respiratory flow as occurs during exercise causes a shift in the O3 dose 

distribution, allowing O3 to penetrate deeper into the lung, increasing the potential for damage to 

bronchiolar and alveolar tissues. 

More recently, Ultman et al. (2004) measured O3 uptake using the bolus technique in 

60 young heathy nonsmoking adults (32 M, 28 F). Bolus were inspired at a rate of 1 L/s, 

equivalent to a moderate exercise rate with a minute ventilation of 30 L/min. Figure 4-2 

illustrates the O3 uptake fraction as a function of VP in a representative subject. Anatomic dead 

space was measured in 47 of the subjects (24 M, 23 F). In these subjects, the volume at which 

50% of an inhaled bolus was absorbed (VP50%) was correlated with VD (r = 0.57, p < 0.001) and 

the volume of the lower conducting airways, i.e., VD minus the volume of the upper airways, 

(r = 0.65, p = 0.001). The better correlation found by subtracting off the upper airways from VD 

can be explained by the fact that very little O3 is absorbed in the upper airways during oral 

breathing at 1 L/s. Both VP50% and lower airways volume were greater in males than females. 

These findings suggest that in females the smaller airways, and associated larger surface-tovolume ratio, enhance local O3 uptake and cause reduced penetration of O3 into the distal lung. 

It is not clear from these findings, however, if the actual anatomical location of VP50% differed 

between males and females. 

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A few studies have measured the effect of a continuous pollutant exposure on O3 bolus 

uptake. Asplund et al. (1996) randomly exposed young healthy adults (8 M, 3 F) for 2 h 

[presumably at rest] to 0.0 (air), 0.12, or 0.36 ppm O3 on 3 separate occasions separated by at 

least 1-wk. Ozone bolus uptake was measured preexposure and subsequently at 30 min intervals 

during the exposure. Ozone uptake over the VP range of 70 to 120 mL increased after the air 

exposure (0.045, absolute change in absorbed fraction), decreased slightly after the 0.12 ppm O3 

exposure (!0.005), and decreased more substantially following the 0.36 ppm O3 exposure 

(!0.03). For clarification, these absolute changes in uptake due to filtered air or O3 exposures 

are increases or decreases from an average uptake of ~0.70 over the VP range from 70 to 

120 mL (Hu et al., 1994). Relative to uptake during the air exposure, Asplund et al. (1996) 

found O3 bolus uptake was significantly decreased by 30 min of the 0.12 and 0.36 ppm O3 

exposures and remained significantly decreased for the duration of these exposures. 

Using a similar protocol, Rigas et al. (1997) randomly exposed young healthy adults (6 M, 

6 F) for 2 h at rest to filtered air, 0.36 ppm NO2, 0.75 ppm NO2, 0.36 ppm SO2, or 0.36 ppm O3. 

Ozone bolus uptake (VP range of 70 to 120 mL ) was measured preexposure and every 30 min 

during the exposures. The results of an F test indicated that exposure duration (30-, 60-, 90-, 

120-min) was not a significant factor, but treatment (NO2, SO2, etc.) was (p < 0.01). Ozone 

bolus uptake was increased by 30 min during the NO2 and SO2 exposures and decreased during 

the O3 exposure. The authors suggested that there may be increased production of an O3-reactive 

substrate in the ELF due to airway inflammation. During NO2 and SO2 exposures the substrate 

was not depleted by these gases and so could react with the O3 bolus. During O3 exposure the 

substrate was depleted, causing the fractional absorption of the O3 bolus to decrease. 

4.2.2 General Uptake Studies 

Ultman and colleagues have recently completed some general uptake studies to determine 

the ratio of O3 uptake to the quantity of O3 inhaled. Uptake efficiency was determined at 

exposures of 0.2 or 0.4 ppm O3 while exercising at a minute volume of approximately 20 L/min 

for 60 min or 40 L/min for 30 min in both men and women (Rigas et al., 2000). Uptake 

efficiency ranged from 0.56 to 0.98 and had a statistically significant but weak dependence on 

4-7
concentration, minute volume, and exposure time. Intersubject differences had the largest 

influence on uptake efficiency, resulting in a variation of approximately 10%. As the quantity 

of O3 retained by the RT is equal to uptake efficiency times the quantity of O3 inhaled, relatively 

large changes in concentration, minute volume, or exposure time may result in relatively large 

changes in the amount of O3 retained by the RT or absorbed locally. The authors concluded that 

for exposure times <2 h, inhaled dose (product of O3 concentration, exposure duration, and 

minute ventilation) is a reasonable predictor of actual uptake as long as there are fixed 

concentrations of O3 and fixed levels of exercise. More importantly, similarly exposed 

individuals vary in the amount of actual dose received. 

Santiago et al. (2001) studied the effects of airflow rate (3 to 15 L/min) and O3 

concentration (0.1, 0.2, or 0.4 ppm) on O3 uptake in nasal cavities of males and females. 

As would be expected, uptake efficiency in the nose was inversely related to the flowrate and the 

concentration of O3 in the inlet air. They computed a gas-phase diffusion resistance of <24% of 

overall diffusion resistance, which suggested to them that simultaneously occurring diffusion 

and chemical reactions in the mucous layer were the limiting factors in O3 uptake. Difference 

in O3 uptake ranged from 0.63 to 0.97 at flowrates of 3 L/min and 0.25 to 0.50 at 15 L/min. 

The small effects of flowrate and concentration on uptake efficiency were statistically 

significant, but intersubject differences accounted for approximately half of the total variation 

in uptake efficiency. Both these general uptake studies, done at environmentally relevant O3 

concentrations, indicate that interindividual differences in fractional uptake are extremely 

important in O3 dose-response relationships. 

In the research mentioned above, Ultman et al. (2004) also completed continuous exposure 

studies. The same 60 subjects were exposed continuously for 1 h to either clean air or 0.25 ppm 

O3 while exercising at a target minute ventilation of 30 L/min. This is the first study to assess 

ventilatory and dosimetric parameters for an entire hour of exposure. In addition to measuring 

pre-to-post exposure changes in FEV1, they used the peripheral bronchial cross-sectional area 

available for diffusion (inferred from the alveolar slope of CO2 expirograms) as an alternative 

response variable. At a fixed minute ventilation of 30 L/min, the uptake fraction of O3 decreased 

with increasing fB (see Figure 4-3) and increased with increasing VT. The uptake fraction was 

4-8
Figure 4-3. Ozone uptake efficiency as a function of breathing frequency at a minute 

ventilation of 30 L/min. The uptake efficiency was well correlated with 

breathing frequency (r = !0.723, p < 0.001) and tidal volume (not illustrated; 

r = 0.490, p < 0.001). 

Source: From Ultman et al. (2004). 

significantly greater in males (91.4%) than females (87.1%), which is consistent with the 

larger fB and smaller VT of the females than males. There was a small but statistically significant 

reduction in the breath-by-breath uptake of O3 from 90.6% on average for the first 15 min to 

87.3% on average for the last 15 min of exposure, although the biological significance of this 

small change is questionable. Ozone uptake rate correlated with percent changes in individual 

bronchial cross-sectional area but did not correlate with individual FEV1 responses. Neither of 

these parameters correlated with the VP50% determined in the bolus studies mentioned above. 

The authors concluded that the intersubject differences in forced respiratory responses were not 

due to differences in O3 uptake. However, these data did partially support the hypothesis that 

changes in cross-sectional area available for gas diffusion are related to overall O3 retention. 

4-9
Plopper et al. (1998) examined the relationship between O3 dose and epithelial injury in 

Rhesus monkeys. Using 18O content in lung tissues, the respiratory bronchioles were confirmed 

as the site receiving the greatest O3 dose (mass 18O per dry lung weight). Furthermore, the 

greatest cellular injury occurred in the vicinity of the respiratory bronchioles and was dependent 

on the delivered O3 dose to these tissues. After a 2 h exposure, the antioxidant glutathione 

(GSH) was increased in the proximal intrapulmonary bronchus after 0.4 ppm O3 and decreased 

in the respiratory bronchiole after 1.0 ppm O3. Perhaps an adaptive response, chronic O3 

exposure leads to increased GSH levels in distal bronchioles of both rats and monkeys relative to 

GSH levels in filtered air-exposed animals (Duan et al., 1996). 

4.2.3 Dosimetry Modeling 

When all of the animal and human in vivo O3 uptake efficiency data are compared, there is 

a good degree of consistency across data sets, which raises the level of confidence with which 

these data sets can be used to support dosimetric model formulations. Models predict that the 

net O3 dose (O3 flux to air-liquid interface) gradually decreases distally from the trachea toward 

the end of the TB and then rapidly decreases in the pulmonary region. However, the tissue 

dose (O3 flux to liquid-tissue interface) is low in the trachea, increases to a maximum in the 

terminal bronchioles and the first generation of the pulmonary region, and then decreases rapidly 

distally into the pulmonary region. The increased VT and flow, associated with exercise in 

humans or CO2-stimulated ventilation increases in rats, shifts O3 dose further into the periphery 

of the lung, causing a disproportionate increase in distal lung dose. 

Table AX4-2 in the annex presents a summary of new theoretical studies of the uptake 

of O3 by the RTs (or regions) of humans and laboratory animals that have become available 

since the 1996 review. These studies are briefly described below. Virtually all of these models 

have assumed that the reaction rate of O3 in the liquid lining layer and in tissues is quasi firstorder with respect to O3 concentration. However, there is considerable discrepancy between rate 

constants used in models. For instance, Hu et al. (1994) and Bush et al. (2001) estimate a 

reaction rate constant that is more than 1000 times as large as that used by Cohen-Hubal et al. 

(1996). Both the uptake of O3 at the gas-liquid interface and the fraction of O3 (or its reaction 

4-10
products) that reach epithelial cells are sensitive to the value of these reaction rate constants. 

A large disparity in rate constants between studies illustrates limitations in our current state of 

knowledge and affects the interpretation of O3 model predictions. 

Overton and Graham (1995) created a rat model combining multiple path anatomic models 

and one-dimensional convection-dispersion equations, which simulates transport and uptake 

of O3 in airways and airspaces of the modeled TB region. Predictions from this model 

realistically detail O3 transport and uptake of different but morphologically equivalent sites. 

Modeled lung tissue doses show variation of O3 dose among anatomically equivalent ventilatory 

units as a function of path length from the trachea with shorter paths generally showing greater 

doses (Overton and Graham, 1995). This conflicts with Schelegle et al. (2001), who exposed 

rats to 1 ppm O3 for 8 h and found that the terminal bronchioles supplied by short and long paths 

had similar epithelial injury. Interestingly, in rats with a C-fiber conduction block to prevent O3­

induced rapid shallow breathing, it was the long path terminal bronchioles that received the 

greatest epithelial injury. Overall, O3-induced rapid shallow breathing appears to protect the 

large conducting airways while producing a more even distribution of injury to the terminal 

bronchioles (Joad et al., 2000; Schelegle et al., 2001). Postlethwait et al. (2000) have also 

identified the conducting airways as a primary site of acute O3-induced cell injury. Such data 

must be considered when developing models that attempt to predict site-specific locations of O3­

induced injury. The early models computed relationships between delivered regional dose and 

response with the assumption that O3 was the active agent responsible for injury. It is now 

known that reactive intermediates such as hydrohydroxyperoxides and aldehydes are important 

agents mediating the response to O3 (further discussed in Section 5.3.1). Thus, models must 

consider O3 reaction/diffusion in the ELF and ELF-derived reactions products. 

Using computational fluid dynamics (CFD), Cohen-Hubal et al. (1996) modeled the effect 

of the mucus layer thickness in the nasal passage of a rat. Predictions of overall uptake were 

within the range of measured uptake. Predicted regional O3 flux was correlated with measured 

cell proliferation for the CFD simulation that incorporated two regions, each with a different 

mucus thickness. 

4-11
With a RT dosimetry model, Overton et al. (1996) investigated the sensitivity of uptake 

efficiency, proximal alveolar region (PAR) dose (O3 mass per unit surface area per unit time), 

and PAR dose ratio to TB region volume (VTB) and TB region expansion in humans and rats. 

The PAR was defined as the first generation distal to terminal bronchioles and the PAR dose 

ratio was defined as the ratio of a rat’s predicted PAR dose to a human’s predicted PAR dose. 

This ratio relates human and rat exposure concentrations so that both species receive the same 

PAR dose. In rats, the PAR is a region of major damage from O3. For each species, three values 

of VTB were used: a mean value from the literature and the mean ± twice the SD. For both the 

rat and human simulations, there were several general findings: (1) uptake efficiency and 

PAR dose both increased with decreasing VTB, e.g., using the highest TB region mass transfer 

coefficient (kTB), the PAR dose for VTB – 2SD was five times greater than the PAR dose 

for VTB + 2SD, (2) uptake efficiency and PAR dose both decreased with TB expansion relative 

to no expansion, (3) PAR dose increased with tidal volume, (4) PAR dose increased with 

decreasing kTB, and (5) uptake efficiency increased with kTB. 

Bush et al. (2001) modified their single-path model (Bush et al., 1996b) so that simulations 

would coincide with experimental uptake efficiency data for O3 and Cl2 during oral and nasal 

breathing. Relative to their original model, the Bush et al. (2001) model added lung expansion 

and modified the mass transfer coefficients for both the gas-phase (kg) and the liquid-phase (kl

). 

Consistent with Overton et al. (1996), considering expansion of the TB airways reduced uptake 

efficiency versus no expansion. As very little inhaled O3 reaches the peripheral lung, it was not 

surprising that alveolar expansion had minimal affect on uptake efficiency. Ignoring the O3 

reaction rate constant (kr), the simulations for O3 and Cl2 were nearly the same since the gasphase diffusion coefficients of O3 and Cl2 are similar. But for a given VP the TB airways of the 

lung, experimental bolus uptake are always less for O3 than for Cl2. The authors surmised that 

the difference between the uptake for these gases could be explained adequately based solely on 

the diffusive resistance of O3 in airways surface fluid (modeled by kr

). Qualitatively, model 

simulations also agreed well with the experimental data of Gerrity et al. (1995). 

4-12
Age- and gender-specific differences in both regional and systemic uptake in humans was 

modeled using a physiologically-based pharmacokinetic (PBPK ) approach (Sarangapani et al., 

2003). The model estimated that regional (URT, TB, pulmonary) extraction efficiency of O3 is 

relatively insensitive to age and gender. 

A recent attempt was made (Mudway and Kelly, 2004) to model O3 dose-inflammatory 

response using a meta-analysis of 23 exposures in published human chamber studies. The O3 

concentrations ranged from 0.08 to 0.6 ppm and the exposure durations ranged from 60 to 

396 min. The analysis showed linear relationships between O3 dose and neutrophilia in 

bronchoalveolar lavage fluid (BALF). Linear relationships were also observed between O3 dose 

and protein leakage into BALF, which suggested to the authors that a large-scale study could 

determine a possible O3 threshold level for these inflammatory responses. These recent findings 

seem consistent with the linear relationship between O3 dose to pulmonary tissues normalized 

for body weight and lavage fluid protein in rats, guinea pigs, and rabbits (Miller et al., 1988). 

4.2.4 Summary and Conclusions - Dosimetry 

Ozone is a highly reactive gas and powerful oxidant with a short half-life. Uptake occurs 

in mucous membranes of the RT where O3 reacts with components of the ELF. Uptake 

efficiency is chemical-reaction dependent and the reaction products (hydrogen peroxide, 

aldehydes, and hydroxyhydroperoxides) created by ozonolysis of polyunsaturated fatty acids 

mediate O3 toxicity. The 1996 O3 AQCD reported that uptake of O3 in rats is about 0.50 and in 

humans at rest is about 0.8 to 0.95. In humans, about 0.07 of the O3 is removed in the 

larynx/trachea, about 0.50 in the head, and about 0.43 in the lungs, where the primary site of 

damage was believed to be the CAR. Increasing flow shifted O3 uptake distally toward smaller 

airways of the lung. Studies in humans showed that increasing minute ventilation with exercise 

(by increasing both breathing frequency and tidal volume) causes only a small decrease in 

uptake efficiency by the total RT. The nasal passages appeared to absorb more O3 than the oral 

passages. Comparing BAL cells, a 0.4 ppm exposure in exercising humans showed 4 to 5 times 

the retained dose of O3 relative to rats exposed at rest to the same concentration. 

4-13
Most new research on O3 uptake has been performed in humans. Bolus-response studies 

demonstrated that a previous continuous exposure to O3 decreases the absorption of a bolus 

of O3, probably due to depletion of compounds able to absorb O3. Continuous exposure to NO2 

and SO2 increased absorption of a bolus of O3. These data are of some relevance to 

environmental exposures where humans may receive differing concentrations of O3 depending 

on time of day. Verifying prior work, the bolus-response method was used to demonstrate 

that O3 bolus uptake is sensitive to the mode of breathing and to the airflow rate. As flow is 

increased from 150 to 1000 mL/s, O3 boli penetrated deeper into the lung and penetration was 

further increased by oral versus nasal breathing. This suggests that the switch from nasal to oral 

breathing coupled with increases in respiratory flow as occurs during exercise causes a shift in 

regional O3 dose deeper into the lung, increasing the potential of damage to bronchiolar and 

alveolar tissues. The finding that O3 uptake is inversely related to airflow also agrees with 

earlier animal studies. 

New general uptake study data demonstrate that exercising men and women receiving 

0.2 or 0.4 ppm O3 at 20 L/min for 60 min or 40 L/min for 30 min absorb 0.56 to 0.98. The 

absorbed fraction was affected only by large changes in concentration, minute volume, and 

exposure time. This suggests that for exposure times <2 h, inhaled dose (i.e., product of O3 

concentration, minute ventilation, and exposure duration) is a reasonable predictor of actual O3 

dose as long as the O3 exposure concentration and level of exercise are relatively constant. 

However, individuals exposed to similar concentrations vary considerably in the amount of 

actual dose received. This intersubject variability has also been demonstrated in a study of O3 

uptake in nasal cavities of men and women. The absorbed fraction in the nose was inversely 

related to the flowrate and the concentration of O3, suggesting that both gas phase diffusion and 

chemical reactions in the mucous layer were limiting O3 uptake. 

The consistency of uptake data generated in animal and human studies allow a high level 

of confidence in their use in dosimetry modeling. Early models predicted that net O3 dose to 

ELF and tissue gradually decreases distally from the trachea toward the end of the TB and then 

rapidly decreases in the pulmonary region. Exercise-induced or CO2-stimulated increases in VT 

and flow, shift O3 dose further into the periphery of the lung, causing a disproportionate increase 

4-14
in distal lung dose. Localized damage to lung tissue has been modeled showing variation of O3 

dose among anatomically equivalent ventilatory units as a function of path length from the 

trachea with shorter paths showing greater damage. 

New models have produced some refinements of earlier models such as: (1) the use of 

mucus resistance and thickness in describing O3 dosimetry and determining the patterns 

of O3-induced lesions; (2) the shape of the dose versus generation plot along any path from the 

trachea to alveoli is independent of path, with the tissue dose decreasing with increasing 

generation index; (3) simulations sensitive to conducting airway volume but relatively 

insensitive to characteristics of the respiratory airspace; (4) the importance of TB region 

expansion; (5) the importance of dose received in the PAR both inter-individual differences and 

extrapolations based on dose; and (6) revaluation of mass transfer coefficients for conducting 

airways. Additionally, more recent data indicate that the primary site of acute cell injury occurs 

in the conducting airways and that reactive intermediates in the ELF, rather than O3 itself, are 

responsible for pulmonary injury. These data must be considered when developing new models. 

4.3 SPECIES HOMOLOGY, SENSITIVITY, AND ANIMAL-TO-HUMAN 

EXTRAPOLATION 

Basic similarities exist across human and other animals species with regard to basic 

anatomy, physiology, biochemistry, cell biology, and disease processes. However, there are 

obviously some species differences that have the potential to affect both the patterns of O3 

uptake in the RT as well as responses. For instance, primates are oronasal breathers with a 

dichotomous branching lung structure, whereas, rodents are obligate nasal breathers with a 

monopodial branching lung structure (Miller et al., 1993). Even when comparing nasal 

breathing, differences in the nasal structure between primates and rodents can affect both the site 

and amount of gaseous uptake in this region (DeSesso, 1993; Morgan et al., 1989). Cellular 

profiles also differ between species as a function of location in the RT (Miller et al., 1993; 

Plopper et al., 1989; Stone et al., 1992). 

4-15
The homology as it exists creates similarities in acute O3-induced effects, especially in the 

RT and in lung defense mechanisms. Rodents appear to have a slightly higher tachypneic 

response to O3, which is clearly concentration-dependent in most species and shows parallel 

exacerbation when hyperventilation (e.g., exercise or CO2) is superimposed. What is not known 

is whether this is evidence of pulmonary irritant sensitivity, perhaps as a prelude to toxicity, or 

whether tachypnea is a defensive action taken by the respiratory system to minimize distal lung 

O3 deposition. Airway or lung resistance in humans is not affected appreciably by acute 

exposure to O3, except under conditions of heavy exercise; animals appear to need high-level 

exposures or special preparations that bypass nasal scrubbing. Dynamic lung compliance (Cdyn) 

has been shown to have small magnitude decreases in response to O3 in some studies across 

species, but it is thought that these changes are of little biological significance for ambient 

exposures. Spirometric changes, the hallmark of O3 response in humans, occur in rats, but to a 

lesser degree. It is unclear, however, the degree to which anesthesia (rat) and the comparability 

of hyperventilation induced by CO2 (rat) or exercise (human) may influence this difference in 

responsiveness. Collectively, the acute functional response of laboratory animals to O3 appears 

quite homologous to that of the human. 

When humans are exposed to O3 repeatedly for several consecutive days, lung function 

decrements subside, and normal spirometric parameters are regained (see Section 6.6). This 

phenomenon of functional attenuation also has been demonstrated in rats, not only in terms of 

spirometry, but also in terms of the classic tachypneic ventilatory response. Full or partial 

attenuation of some BAL parameters also appears to occur in both rats and humans, but exposure 

scenario appears to play a role; other cellular changes do not attenuate (see Section 6.9.4). 

Existing epidemiologic studies provide only suggestive evidence that persistent or progressive 

deterioration in lung function is associated with long-term oxidant-pollutant exposure (see 

Chapter 7). With chronic, repeated exposures to $0.12 ppm O3, however, laboratory animals 

demonstrate changes in lung structure, function, and biochemistry that are indicative of airway 

irritation and inflammation with the possible development of chronic lung disease (U.S. 

Environmental Protection Agency, 1996). Based on the apparent homology of these responses 

4-16
between humans and laboratory animals, animal studies appear to provide a means for assessing 

such chronic health concerns. 

For similarly exposed animals (i.e., same O3 concentration and exposure duration), intraand inter-species differences in pulmonary responses are observed as a function of animal age, 

ventilation, and antioxidant status. Examination of BAL constituents show that the influx of 

inflammatory cells and protein from the serum is influenced by species, but perhaps to less 

extent than by ventilation and antioxidant status. Based on lavage protein levels (mg protein 

per mL lavage fluid), Hatch et al. (1986) reported that guinea pigs were the most responsive 

(to $0.2 ppm); rabbits were the least responsive (2.0 ppm only); and rats, hamsters, and mice 

were intermediate (effects at 0.5 to1.0 ppm). Recognizing that there are differences in O3 doses 

to tissues between species, Miller et al. (1988) examined the relationship protein levels reported 

by Hatch et al. (1986) and predicted pulmonary tissue dose (mass O3 per body weight). Miller 

et al. (1988) found that protein levels in guinea pigs increased more rapidly with tissue dose than 

in rats and rabbits. 

A species’ susceptibility to the effects of O3 exposure may be due, in part, to biochemical 

differences among species. Evidence for this is provided by differences in activity of SD rat and 

rhesus monkey CYP moonoxygenases elicited by O3 exposure (Lee et al., 1998). Additional 

characterization of species- and region-specific CYP enzymes will create a better understanding 

of the differences in response to O3. This will allow more accurate extrapolation from animal 

exposures to human exposures and toxic effects. 

Antioxidant metabolism varies widely among species, which can greatly influence the 

effects of O3 (discussed in greater detail in 5.2.1.3). The guinea pig appears to be the species 

most susceptible to O3. Early studies ranked mice > rats > guinea pigs in order of antioxidant 

responsiveness to O3 challenge. Guinea pigs have been shown to have lower basal levels of 

GSH transferase activity, lower activity of GSH peroxidases, and lower levels of vitamin E 

compared to rats. However, differences in the levels of antioxidants between species and 

regions of the lung do not appear to be the primary factor determining susceptibility to O3­

induced tissue injury (Duan et al., 1993, 1996). Plopper et al. (1998) concluded that in monkeys 

there was a close association between site-specific O3 dose, the degree of epithelial injury, and 

4-17
reduced-glutathione depletion. Within a species, antioxidant defenses against O3 can also vary 

with animal age (Servias et al., 2005) and exposure history (Duan et al., 1996). 

Because cytokine and chemokine responses are so important in an animal’s defense 

against O3 exposure, comparisons of differences in species expression and activity of these 

inflammatory mediators is necessary. Arsalane et al. (1995) compared guinea pig and human 

AM recovered in BALF and subsequently exposed in vitro to 0.1 to 1 ppm for 60 min. 

Measurement of inflammatory cytokines showed a peak at 0.4 ppm in both species. Guinea pig 

AM had an increase in IL-6 and TNFa while human AM had increases in TNFa, IL-1b, IL-6 and 

IL-8. This exposure also caused an increase in mRNA expression for TNFa, IL-1b, IL-6 and 

IL-8 in human cells. At 0.1 ppm exposures, only TNFa secretion was increased. These data 

suggest both qualitatively and quantitatively similar cytokine responses in AM from guinea pigs 

and humans. However, these in vitro AM responses can not be extended directly to an in vivo 

scenario since similar O3 exposures (concentration and duration) do not give the same O3 doses 

in different species (Hatch et al., 1994; Miller et al., 1978). 

Species differences in morphological responses to O3 exposure have been characterized by 

Dormans et al. (1999), as discussed in previous sections. Dormans et al. (1999) continuously 

exposed rats, mice, and male guinea pigs to filtered air, 0.2, or 0.4 ppm O3 for 3, 7, 28, and 

56 days. The animals exposed for 28 days were examined at 3, 7, or 28 days PE. Depending 

on the endpoint studied, the species varied in sensitivity. Greater sensitivity was shown in the 

mouse as determined by biochemical endpoints, persistence of bronchiolar epithelial 

hypertrophy, and recovery time. Guinea pigs were more sensitive in terms of the inflammatory 

response though all three species had increases in the inflammatory response after three days that 

did not decrease with exposure. These data on inflammation are in general agreement with 

Hatch et al., (1986), discussed above. In all species, the longest exposure to the highest O3 

concentration caused increased collagen in ductal septa and large lamellar bodies in Type II 

cells, but that response also occurred in rats and guinea pigs at 0.2 ppm. No fibrosis was seen at 

the shorter exposure times and the authors question whether fibrosis occurs in healthy humans 

after continuous exposure. The authors do not rule out the possibility that some of these 

4-18
differences may be attributable to differences in total inhaled dose or dose actually reaching a 

target site. Overall, the authors rated mice as most susceptible, followed by guinea pigs and rats. 

Comparisons of airway effects in rats, monkeys and ferrets resulting from exposures of 

1.0 ppm O3 for 8 h (Sterner-Kock et al. 2000) demonstrated that monkeys and ferrets had similar 

inflammatory responses and epithelial necrosis. The response of these two species was more 

severe than that seen in rats. These data suggest that ferrets are a good animal model for O3­

induced airway effects due to the similarities in pulmonary structure between primates and 

ferrets. However, the mechanisms of O3 effects at these high concentrations may differ from 

those at more realistic levels. 

A number of species, including nonhuman primates, dogs, cats, rabbits, and rodents, have 

been used to study the effects of O3 exposure on airway bronchoconstriction. A commonly used 

model of bronchospasm utilizes guinea pigs acutely exposed to high O3 concentrations (2 to 

3 ppm) to induce airway hyperreactivity (AHR). As mentioned earlier, the model is helpful for 

determining mechanistic aspects of AHR, but is not really relevant for extrapolation to potential 

airway responses in humans exposed to ambient levels of O3. Additionally, guinea pigs have 

been shown to have AHR in other studies that is very similar to asthmatic humans, but the utility 

of guinea pig data is somewhat limited by their disparity from other animal models. 

The rat is a key species used in O3 toxicological studies, but the rat has both behavioral 

and physiological mechanisms that can lower core temperature in response to acute exposures, 

thus limiting extrapolation of rat data to humans. Iwasaki et al. (1998) evaluated cardiovascular 

and thermoregulatory responses to O3 at exposure of 0.1, 0.3, and 0.5 ppm O3 8 hrs/day for 

4 consecutive days. A dose-dependent disruption of HR and Tco was seen on the first and second 

days of exposure, which then recovered to control values. Watkinson et al. (2003) exposed 

rats to 0.5 ppm O3 and observed this hypothermic response, which included lowered HR, 

lowered Tco, and increased inflammatory components in BALF. The authors suggested that the 

response is an inherent reflexive pattern that can possibly attenuate O3 toxicity in rodents. They 

discuss the cascade of effects created by decreases in Tco, which include: (1) lowered metabolic 

rate, (2) altered enzyme kinetics, (3) altered membrane function, (4) decreased oxygen 

consumption and demand, (5) reductions in minute ventilation, which would act to limit the dose 

4-19
of O3 delivered to the lungs. These effects are concurrent with changes in HR which lead to: 

(1) decreased CO, (2) lowered BP, and (3) decreased tissue perfusion, all of which may lead to 

functional deficits. The hypothermic response has not been observed in humans except at very 

high exposures, which complicates extrapolation of effects in rats to humans. 

The importance of animal studies derives from their utilization in determining cause-effect 

relationships between exposure and health outcome, but the animal data must be integrated with 

epidemiological studies and controlled human clinical studies. Animal studies can corroborate 

both clinical and epidemiology studies and further provide important data that is impossible to 

collect in human studies. Toxic pulmonary and extrapulmonary effects following O3 exposure 

have been well-studied in rodents, nonhuman primates, and a few other species; so, 

extrapolation, both qualitative and quantitative, to human exposures and consequent health 

effects is possible. Quantitative extrapolation, required to determine what specific exposure is 

likely to cause an effect in humans, is theoretically founded on the equivalency of mechanisms 

across species. At the molecular level, O3 acts on the carbon-carbon double bond in 

polyunsaturated fatty acids and on sulfhydryl groups in proteins, both of which are found within 

cell membranes in animals and humans. At higher levels of cellular organization, cells affected 

in animals (e.g., AMs, Type 1 cells) have similar functions in humans, and organ systems (e.g., 

respiratory system) have major interspecies similarities. However, interspecies differences do 

occur and complicate extrapolation. 

Quantitative extrapolation, which involves a combination of dosimetry and species 

sensitivity, still requires more research before it can be fully realized. Knowledge of dosimetric 

animal-to-human extrapolation is more advanced than that of species-sensitivity, but 

extrapolation models have not been completely validated, and therefore, significant uncertainties 

remain. Mathematical modeling of O3 deposition in the lower RT (i.e., from the trachea to 

alveoli) of several animal species and humans shows that the pattern of regional dose is similar, 

but that absolute values differ. In spite of structural and ventilatory differences between species, 

the greatest predicted tissue dose is to the CAR. Even though the CAR of rats has very 

rudimentary respiratory bronchioles, compared to well-developed ones in primates, the CAR of 

both rats and nonhuman primates respond similarly to O3. 

4-20
Experimental measurement of delivered O3 doses estimate that total respiratory uptake is 

~47% in laboratory animals and ~87% in exercising humans, while nasopharyngeal removal is 

~17% in rats and ~40% in humans. The previous O3 AQCD (U.S. Environmental Protection 

Agency, 1996) provided the first quantitative animal-to-human extrapolation of morphological 

changes in the proximal alveolar region using rat and monkey studies. The extrapolation 

predicted that a 9-year-old child would have a 20% or 75% increase in PAR tissue thickness if 

their sensitivity to O3 was equal to that of a rat or monkey, respectively. Adults would have 

15 or 70% increase, suggesting the potential for chronic effects in humans. In spite of the 

significant uncertainties, this extrapolation raises concern about the potential for chronic effects 

in humans 

Experiments using 2 h exposures to 0.4 ppm 18O3 suggested that exercising (15 min 

intervals, rest and exercise at 60 L/min) humans received a 4- to 5-fold higher 18O3 

concentrations in BAL than resting rats (Hatch et al., 1994). That level of exposure increased 

BAL protein and PMNs in humans, while a concentration of 2.0 ppm in rats was necessary for 

similar effects. Caveats in the interpretation of 18O3 studies include: (1) only a very small 

portion of the labeled compound is recoverable to assess incorporation; and (2) if species being 

compared differ in physiocochemical factors controlling mass transfer and downstream O3 

metabolism, it could cause significant differences in the amount of inhaled 18O3 that is detected 

during subsequent tissue analysis. Further, species differences in pulmonary anatomy, 

ventilation, antioxidants, and susceptibility all influence dose, repair processes, and tolerance to 

subsequent O3 exposure. Important differences between exercising humans and resting rats that 

can affect tissue O3 dose include: (1) increased ventilation and O3 delivery with exercise; 

(2) decreased pulmonary ventilation and body temperature during O3 exposure in rats; 

(3) diminished dose received in rats due to their burying their noses in their fur during exposure; 

and (4) increased concentration of antioxidants in ELF in rats compared to humans. These 

antioxidants are important for converting O3 to inactive products before toxicity occurs (Kari 

et al., 1997; Gunnison and Hatch, 1999; Plopper et al., 1998), though this quenching is not 

quantitative. These and possibly other differences between rats and humans suggest that a 

2 ppm exposure in nonexercising rats approximates a 0.4 ppm exposure in exercising humans. 

4-21
Further comparisons of exercising human exposure to 0.1 ppm for 6 hours (Devlin et al., 1991) 

and resting rat exposure to 0.3 ppm show inflammatory and permeability changes in humans but 

not rats. 

4.3.1 Summary and Conclusions: Species Homology, Sensitivity, and 

Animal-to-Human Extrapolation 

Comparisons of acute exposures in rats and humans suggest that, though both species have 

similar qualitative responses to O3 exposure, there are interspecies mechanistic disparities that 

necessitate careful comparisons of dose-response relationships. There is no perfect nonhuman 

species with which to model O3 toxicity. All have limitations that must be considered when 

attempting to extrapolate to human exposures. Awareness of these limitations, even at the level 

of subtle strain differences within a test species, is extremely important. The currently available 

data suggest that LOELs in resting rats are approximately 4- to 5-fold higher than for exercising 

humans for toxicological endpoints including BAL protein and BAL PMNs. Studies comparing 

species-specific differences in O3-induced effects showed that guinea pigs were the most 

susceptible, rabbits the least susceptible, and rodents intermediate in susceptibility. The recent 

work being done utilizing various mouse strains with differing sensitivities to O3 will help us to 

understand the extremely complex inter-individual differences in human sensitivity to O3. 

4-22
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5. TOXICOLOGICAL EFFECTS OF OZONE AND

RELATED PHOTOCHEMICAL OXIDANTS IN

LABORATORY ANIMALS AND IN VITRO

TEST SYSTEMS

5.1 INTRODUCTION 

A wide range of effects of ozone (O3) has been demonstrated in laboratory animals. The 

major research findings are that environmentally relevant levels of O3 cause lung inflammation; 

decreases in host defenses against infectious lung disease; acute changes in lung function, 

structure, and metabolism; chronic lung disease, some elements of which are irreversible; and 

systemic effects on target organs (e.g., brain, heart, liver, immune system) distant from the lung. 

The research also has served to expand the understanding of mechanisms of O3 toxicity and the 

relationships between concentration and duration of exposure. 

The framework for presenting the health effects of O3 in animals begins with a presentation 

of respiratory tract effects, followed by systemic effects, and then interactions of O3 with other 

common co-occurring pollutants. The information discussed in this chapter is founded on a very 

wide body of literature on studies in laboratory animals and on in vitro test systems of animal 

cell lines and organ systems that may mimic responses in intact animals. The direct effects of O3 

in humans are discussed in the following chapter (Chapter 6). 

This chapter is not intended to be a compendium of all that is known about toxicologic 

effects of O3; rather, it is an update of the toxicology chapter from the last previous O3 criteria 

document (U.S. Environmental Protection Agency, 1996), or 1996 O3 AQCD, and other reviews 

of the earlier published literature. The historical O3 literature is very briefly summarized in an 

opening paragraph of each section or subsection. That paragraph is intended as a very concise 

overview of previous work, and the reader is referred to the 1996 O3 AQCD for more detailed 

discussion of the literature prior to the early 1990’s. Each section then continues with brief 

discussions of the key new studies (or somewhat older studies that were not included in the 

1996 O3 AQCD). Longer discussions of new studies are included where warranted. Sections are 

ended with comparisons of data from the previous AQCD with new data, and basic conclusions 

are drawn. Summaries of new studies and results are provided in tables in Annex AX5. 

5-1
Except for nitrogen dioxide (NO2), the subject of another criteria document (U.S. 

Environmental Protection Agency, 1993), there is very little relevant information on other 

photochemical oxidants in the published literature. What is known about the effects of these 

other oxidants is also summarized briefly in this chapter. 

5.2 RESPIRATORY TRACT EFFECTS OF OZONE 

5.2.1 Biochemical Effects 

Biochemically detected effects of O3 are integrally involved in effects on both structure 

and function (respiratory and nonrespiratory) of the respiratory tract. Changes in xenobiotic 

metabolism, antioxidant metabolism and oxygen consumption, lipids and arachidonic acid 

metabolism, and collagen metabolism are all observed with O3 exposure, though the mechanisms 

and associations are not fully understood. 

5.2.1.1 Cellular Targets of Ozone Interaction 

Ozone has the potential to interact with a wide range of different cellular components that 

include polyunsaturated fatty acids (PUFAs); some protein amino acid residues; and some 

low-molecular-weight compounds that include glutathione (GSH), urate, vitamins C and E, and 

free amino acids. Early work demonstrated that O3, being a highly reactive compound, does not 

penetrate much beyond the epithelial lining fluid (ELF); and reaction/diffusion analyses suggest 

that O3, at environmentally-relevant concentrations, diffuses no more than 0.1 to 0.2 μm into the 

ELF. However, Miller (1995) points out that throughout the respiratory tract, the ELF varies in 

thickness and that the distal conducting airways may have only a patchy lining layer which 

allows O3 to react directly with cell membranes. Various models utilizing differing reaction 

kinetics and rate constants do not agree as to whether or not inhaled O3 can penetrate the ELF 

and reach the epithelial cells. Ozone-induced cell damage results, in part, from its reactions with 

PUFAs to form stable but less reactive ozonide, aldehyde, and hydroperoxide reaction products. 

These reaction products (Criegee ozonides and hydroxyhydroperoxides) may act as signal 

transduction molecules involved in signaling of cellular responses such as inflammation and thus 

mediate O3 toxicity. These reactions are summarized in Figure 5-1. Studies published since the 

1996 AQCD are listed in Table AX5-1. 

5-2
Figure 5-1. Schematic overview of ozone interaction with epithelial lining fluid and lung 

cells. It should be noted that not all secondary reaction products are shown. 

Frampton et al. (1999) demonstrated the ozonation of PUFA to form nonanal and hexanal 

in rat bronchioalveolar lavage (BAL) after exposures to 0.22 ppm O3 for 4 h with exercise. 

Increases in nonanal were not accompanied by significant changes in lung function, in epithelial 

permeability, or in airway inflammation. Hexanal levels did not increase significantly and levels 

of both aldehydes returned to baseline by 18 h postexposure (PE). Pryor et al. (1996) exposed 

rats to 0.5 to 10 ppm O3 both with and without 5% CO2 to measure the amount of aldehyde 

generated in BAL and also the rate of disappearance of aldehydes from the ELF following the O3 

exposure. Ozone exposure with CO2 increased the tidal volume and the yield of aldehydes, with 

a maximal aldehyde yield at 2.5 ppm for 1 h. Absolute yields were impossible to ascertain in 

this system, because absorption of O3 is unknown and aldehyde recovery is not complete due to 

loss of aldehyde by volatization and by diffusion into underlying tissue. The data showed that at 

0.5 ppm O3 with 5% CO2, levels of hexanal and nonanal increased at 30 min, decreased slightly 

from that levels at 60 min, was maximal at 90 min and then dropped to 60 min levels at 120 min. 

Heptanal levels did not change appreciably during this time course. Levels of these aldehydes 

were dependent on a dynamic relationship between their production and disappearance from the 

ELF. The authors stated that O3 is the limiting reagent in this process because the amount of 

PUFA far exceeds the amount of O3 on a molar basis. Because of the limitations of measuring 

aldehydes in this study protocol, the results are not useful for quantitative dosimetry; however, 

5-3
the authors suggest that the study does serve to demonstrate the use of aldehydes as biomarkers 

of O3 exposure, as nonanal is produced in an O3-specific pathway. 

Postlethwait et al. (1998) utilized three biologically relevant models (isolated epithelial 

lining fluid, intact lung, and liposome suspensions) to determine the O3-induced production of 

heptanal, nonanal and hexanal in an attempt to estimate formation of lipid-derived bioactive 

compounds. Exposures used were 0.25 to 1.0 ppm for 30 to 60 min. The results suggest that 

PUFAs directly react with O3 and the amount of bioactive lipids produced is inversely related to 

ascorbic acid (AA) availability. The authors caution that there are limitations to the use of 

measurements of these reaction products in determining O3 dose-response relationships due to 

the heterogenous nature of O3 reactions in the epithelial lining fluid. Connor et al. (2004) have 

recently examined the reactive absorption of O3 (0.3 to 1.1 ppm for 1 to 2 h) within ELF using 

interfacial films composed of dipalmitoylglycero-3-phosphocholine (DPPC) and rat lung lavage 

fluid. The films reduced O3-reactive absorption by antioxidants. Further experiments using a 

human lung fibroblast cell line exposed to O3 demonstrated that AA produced cell injury, that 

high levels of O3 and AA were needed to induce cell injury, and that DPPC films reduced the 

amount of cell injury. Based on these data, the authors suggest that O3 reactions with ELF 

substrates cause cell injury, that films of active, saturated phospholipids reduce the local dose 

of O3-derived reaction products, and that these interfacial phospholipids modulate the 

distribution of inhaled O3 and the extent of site-specific cell injury. 

Recent studies have examined the formation of ozonation products such as 

4-hydroxynonenal (HNE), a toxic aldehyde that reacts with cysteine, histamine, and lysine 

amino acid residues and creates protein adducts. Hamilton et al. (1998) demonstrated (see 

Chapter 6) using human alveolar macrophages (AMs) exposed to 0.4 ppm O3 for 1 h that 

exposure caused apoptosis, an increase in a 32-kDa protein adduct, and an increase in ferritin 

and a 72-kDa heat shock protein. By exposing AM to HNE in vitro, all of these effects are 

replicated, which the authors interpret to mean that creation of protein adducts and apoptotic cell 

death are cellular toxic effects of acute O3 exposure that are mediated, at least in part, by HNE. 

These recent reports, combined with observations reported in the previous O3 AQCD (U.S. 

Environmental Protection Agency, 1996), suggest that interactions of O3 with cellular 

components and ELF generate toxic ozonation products and mediate toxic effects through these 

products. 

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5.2.1.2 Monooxygenases 

Both short- and long-term exposures to O3 have been shown to enhance lung xenobiotic 

metabolism, possibly as a result of changes in the number and function of bronchiolar epithelial 

Clara cells and alveolar epithelial Type II cells. Studies of the effects of O3 on lung 

monooxygenases are listed in Table AX5-2. Early studies showed that exposure to O3 increased 

CYP 2B1 (the major CYP isoform in rat lung) content and activity in rat lung. Ozone exposures 

also caused hypertrophy and hyperplasia of CYP 2B1-immunoreactive Clara cells. Comparisons 

of rat and rhesus monkey CYP isoforms demonstrated species-specific and region-specific (e.g., 

trachea, parenchyma) differences in the activities of P450 isoforms (Lee et al., 1998). 

Watt et al. (1998) found that, with both acute (8 h, 1 ppm) and chronic (90 days, 1 ppm) 

exposures in rat, 1 ppm O3 increased CYP 2E1 in a region-specific manner. Paige et al. (2000a) 

showed that a long-term exposure (0.8 ppm, 8 h/day for 90 days) increased the activity of CYP 

2B in distal lung but not in trachea or intrapulmonary airways. Studies have focused on P450 

gene expression to examine possible genetic mechanisms that may explain differential O3­

sensitivity (Mango et al., 1998). Mice (129 strain) deficient in Clara cell secretory protein 

(CCSP-/-), which are oxidant-sensitive, were exposed to 1 ppm O3 for 2 hours. The CCSP null 

mice demonstrated increases in interleukin-6 (IL-6) and metallothionein (MT) mRNA that 

preceded decreases in Clara cell CYP 2F2 mRNA (normally expressed at high levels in mouse 

lung) levels. In 129 strain wild-type (WT) mice, RNA levels changed similarly, but to a lesser 

degree. These data suggest a protective role against oxidant damage for CCSP, and further, that 

genetic susceptibility to oxidant stress may be mediated, in part, by the gene coding for CCSP. 

5.2.1.3 Antioxidants, Antioxidant Metabolism, and Mitochondrial Oxygen Consumption 

Ozone also undergoes reactions with AA, reduced glutathione (GSH), and uric acid (UA), 

all antioxidants present in ELF (see Figure 5-2, A). In vivo experiments have shown that 

reactions with O3 occur preferentially with antioxidants compared to proteins and lipids also 

present in ELF. This is a protective interaction, but even with environmentally relevant 

exposures to O3, the reactivity of O3 is not quantitatively quenched. Antioxidants offer some 

protection from O3 exposure but often are not maintained at concentrations sufficient to fully 

protect the lung. Thus, O3-induced cell injury occurs in both the lower and upper respiratory 

tract. Early work has shown that acute (1 week) exposures to <1 ppm O3 increase antioxidant 

5-5
Figure 5-2. The major cellular targets and proposed mechanisms of ozone toxicity in the 

lung. Abbreviations: GSH - reduced glutathione; AA - ascorbate; UA - urate; 

"-Toc - alpha-tocopherol; GSH-Px - glutathione peroxidase; EC-SOD -

extracelluar superoxide dismutase; T.I.C. - Type I cells; T.II.C. - Type II cells; 

Neut - neutrophils; Lymp - lymphocytes; Epi. - epithelium; Inst. - interstitum; 

Endo - endothelium; M.C. - monocyte; C.E.C. - ciliated epithelial cells; A.M. -

alveolar macrophage; B.E.C. - bronchial epithelial cells; ROS - reactive oxygen 

species. 

Source: Mudway and Kelly (2000), with permission. 

metabolism, including levels of cytosolic enzymes glucose-6-phosphate dehydrogenase (G6PD), 

6-phosphogluconate dehydrogenase (6PGD), glutathione reductase (GR), and glutathione 

peroxidase (GSHPx). Reexposure after a recovery period causes increases equivalent to firsttime exposures; thus, previous exposure does not appear to be protective. 

Increases in enzyme activity appear to increase as a function of age, suggesting that O3 

exposure can cause greater lung injury in the older animal. This has been attributed to 

differences in dose reaching lung target sites, differing base levels of antioxidants and 

antioxidant enzymes, and differences in cellular sensitivity. Species differences exist in 

antioxidant metabolism, with guinea pigs being very sensitive to O3 due to their diminished 

increases in antioxidants and antioxidant enzymes. Chronic exposures of rats to urban patterns 

5-6
of O3 (daily peaks of 0.25 ppm) caused increases in GSHPx and GR, but not superoxide 

dismutase (SOD). The enzyme changes could be accounted for by differences in the steady-state 

cell population or in cellular antioxidant capacity. More recent studies examining antioxidants 

and O3 exposure are listed in Table AX5-3. 

Ozone induced both site- and cell-specific changes in copper-zinc (Cu-Zn) and manganese 

(Mn) SOD in rats exposed to 1.0 ppm O3 for up to 3 months (Weller et al., 1997). CuZnSOD 

labeling was decreased in epithelial cells in airways and parenchyma. Mn-SOD labeling was 

increased in both AM and epithelial type II cells of the centriacinar region (CAR), which the 

authors suggest may allow these cells to tolerate further O3 exposure. This work is in agreement 

with earlier work suggesting a role of SOD in protection of cells against oxidative stress. 

Freed et al. (1999) evaluated the role of antioxidants in O3-induced oxidant stress in dogs 

(exposed to 0.2 ppm for 6 h) by inhibiting the antioxidant transport using probenecid (an aniontransport inhibitor). Blocking antioxidant transport caused heterogeneously distributed increases 

in peripheral airway resistance (Raw) and reactivity, supporting the hypothesis that in the lung 

periphery, endogenous antioxidants moderate the effects of O3 and that this exposure is a 

subthreshold stimulus for producing effects on peripheral Raw and reactivity in dogs. The 

authors further found that treatment with probenecid also inhibited O3-induced neutrophilic 

inflammation, providing evidence for a dissociation between airway function and inflammation. 

This suggests that O3-induced inflammation and airway hyperreactivity (AHR) are independent 

phenomena operating through multiple mechanistic pathways. 

Mudway and Kelly (1998) modeled the interactions of O3 with ELF antioxidants using a 

continually mixed, interfacial exposure set up with O3 concentrations ranging from 0.1 to 

1.5 ppm for durations ranging from 30 to 720 min. Uric acid was ranked the most O3-reactive, 

AA the second most reactive, and GSH the least reactive. Thus, they concluded that GSH is not 

an important substrate for O3, while UA appeared to be the most important reactive substrate, 

which confers protection from O3 by removing it from inhaled air and limiting the amount that 

reaches the distal lung. By providing a substrate for O3 reactions in the ELF, UA effectively 

reduces the diffusion resistance of O3 (see Bush et al., 2001) in the tracheobronchial airways and 

thus may serve to limit the amount of O3 reaching the distal lung. The authors acknowledge 

limitations in extrapolating these data to in vivo O3 exposures due to the absence of surfactant 

lipids and airway mucus in this in vitro model. 

5-7
5.2.1.4 Lipid Metabolism and Content of the Lung 

One of the major postulated molecular mechanisms of action of O3 is peroxidation of 

mono- and polyunsaturated fatty acids and unsaturated neutral lipids in the lung. Because all of 

these lipids appear both in cell membranes and as secretions in the ELF, it is difficult to ascertain 

which lipid pool contributes to the formation of lipid ozonation products. As mentioned, O3 can 

penetrate only a short distance into the ELF; and, therefore, it reacts with epithelial cell 

membranes only in regions of distal lung where ELF is very thin or absent. The inflammatory 

cascade (shown in Figure 5-3) initiated by O3 generates a mix of secondary reactants (e.g., 

aldehydes), which then are likely to oxidize lipids and proteins in cell membranes. 

In both acute and short-term studies, a variety of lung lipid changes occur, including an 

increase in arachidonic acid. Metabolism of arachidonic acid produces a variety of biologically 

active mediators that can, in turn, affect host defenses, lung function, the immune system, and 

other functions. The protein A component of surfactant is also a primary target of O3 interaction. 

During the first few days of O3 exposure, the changes in lung lipid biosynthesis can be accounted 

for by the alveolar epithelial proliferative repair. With longer exposures (e.g., 0.12 ppm for 

90 days) an increase in PUFAs and a decrease in cholesterol-esters are seen, indicative of 

long-term alterations of surfactant lipid composition. 

Several new studies listed in Table AX5-4 examined the effects of O3 exposure on 

phospholipids in lung tissue. Ozonation of PUFAs has been shown to generate other aldehydes 

such as nonanal and hexanal in rats (Pryor et al., 1996; Frampton et al., 1999). These aldehydes 

are short-lived and found to not affect lung function (Frampton et al., 1999). These observations 

suggest that levels of these aldehydes are dependent on a dynamic relationship between their 

production and their disappearance from the ELF. 

Pryor et al. (1995) proposed a cascade mechanism whereby ozonation products cause 

activation of specific lipases, which then trigger the activation of second messenger pathways 

(e.g., phospholipase A2 or phospholipase C). This group (Kafoury et al., 1999) showed that 

exposure of cultured human bronchial epithelial cells to the lipid ozonation product 1-palmitoyl2-(9-oxononanoyl)-sn-glycero-3-phosphocholine elicited release of platelet-activating factor 

(PAF) and PGE2, but not IL-6. The lipid ozonation product 1-hydroxy-1-hydroperoxynonane 

caused release of PAF and IL-6 in these cells, but not PGE2. These results suggest to the authors 

5-8
Figure 5-3. Mechanisms of ozone toxicity. The inflammatory cascade starts by interactions of O3 

with lipids in the epithelial lining fluid and in cell membranes. Lipid ozonation 

reaction products include aldehydes, hydroxyhydroperoxides, and H2O2. These 

initiate lipase activation in epithelial cells resulting in the downstream production of 

phopholipases, arachidonic acid, leukotrienes, and prostanoids. Leukotrienes and 

prostanoids alter lung function. Phospholipases initiate the production of NO, 

inflammatory cytokines, and PAF. NO, IL-6, and IL-8 then participate in the 

recruitment and activation of inflammatory cells (AMs, PMNs, blood monocytes, and 

mast cells), which, in turn, can start a second cycle of inflammatory responses. PAF 

increases epithelial permeability and may have thrombolytic effects. Release of 

substance P by neurons causes altered lung function and increases vascular 

permeability. Production of NOS-1 by these neurons induces airway hyperreactivity. 

5-9
that O3-induced production of lipid ozonation products causes release of proinflammatory 

mediators that then generate an early inflammatory response. 

Very new work (Ballinger et al., 2005) has shown that O3-induced membrane oxidation is 

augmented by antioxidants present in ELF. They utilized a red cell membrane model exposed to 

0.8 ppm O3 for 30 min. The monolayer of cells was intermittently covered by an aqueous film 

consisting of rat bronchioalveolar lavage fluid (BALF) or BALF plus added antioxidants. 

Ascorbic acid and GSH induced dose-dependent oxidative damage to the cell membrane proteins 

and lipids via secondary oxidant formation. The authors concluded that early in O3 exposure, 

ELF antioxidants are high enough to drive reactive absorption of O3 into the ELF and to 

concurrently quench secondary reaction products, thus limiting cell injury. With continued 

exposure, levels of antioxidants decrease such that unreacted O3 and cytotoxic products can 

diffuse to the cell membranes, causing injury. Limitations of this in vitro study include possible 

differences in chemical species and mechanisms compared to in vivo systems. 

Uhlson et al. (2002) reacted O3 with calf lung surfactant which resulted in the production 

of 1-palmitoyl-2-(9'-oxo-nonanoyl)-glycerophosphocholine (16:0a/9-al-GPCho). The biological 

activity of this oxidized phospholipid included: (1) decreased macrophage viability, 

(2) induction of apoptosis in pulmonary epithelial-like A549 cells, (3) and release of IL-8 from 

A549 cells. Exposures to 0.125 ppm O3 for 2 to 4 h in this system were capable of generating 

biologically active phospholipids that were capable of mediating toxic effects of O3. 

In addition to PUFA, cholesterol, the most abundant neutral lipid present in ELF, is also a 

target of O3. Pulfer and Murphy (2004) demonstrated the ozonolysis of cholesterol in an in vitro 

system using BALF isolated from rats that had been exposed to 2.0 ppm O3 for 4 h. Production 

of 5-hydroperoxy-B-homo-6-oxa-cholestan-3,7a-diol, 5$,6$-epoxycholesterol, and 3$-hydroxy5-oxo-5,6-seco-cholestan-6-al was shown. Additionally, both 5$,6$-epoxycholesterol and its 

most abundant metabolite, cholestan-6-oxo-3$,5"-diol, were demonstrated to be cytotoxic to 

16-HBE cells and to inhibit cholesterol synthesis. Studies (Pulfer et al., 2005) in C57BL/6J mice 

exposed to 0.5, 1.0, 2.0, or 3.0 ppm O3 for 3 h demonstrated that these oxysterols were produced 

in vivo also. The authors suggested that this may be an additional mechanism of O3 toxicity. 

Though these oxysterol reaction products have not been fully characterized, they may be 

involved in O3-induced inflammation by disrupting cellular membranes or altering signaling 

5-10
between cells. Similar oxysterols have been implicated in the inflammatory cascade associated 

with atherosclerosis. 

Thus, new work has attempted to elucidate the mechanisms by which reactions of O3 with 

lipids create phospholipids that then mediate downstream toxic effects. It is uncertain whether 

these described changes in lipid content and/or metabolism lead to significant changes in surface 

tension or compliance properties of the lung. 

5.2.1.5 Ozone Interactions with Proteins and Effects on Protein Synthesis 

Epithelial lining fluid contains proteins arising from airway secretions and from blood. 

Ozone can react with four amino acid residues (cysteine, histidine, methionine, and tryptophan) 

and can cause oxidation of functional groups on proteins, including aldehydes, alcohol, amines 

and sulfhydryls. A number of enzymes have been shown to be inhibited by O3, including 

cholinesterase, "1-antiproteinase, and prostaglandin synthetase. Additionally, O3 decreases the 

inhibitory activity of "1-proteinase inhibitor, which is implicated in development of emphysema. 

Surfactant protein A (SP-A) is a target for O3 toxicity by modulation of SP-A self association, 

vesicle aggregation, phospholipid secretion, and stimulation of AM superoxide anion generation 

(see Section 5.2.2.3). Further, O3 is thought to interfere in SP-A’s homeostatic role in surfactant 

release from alveolar Type II cell lamellar bodies and its subsequent uptake by Type II cells 

and AMs. 

Lung collagen, collagen synthesis, and prolyl hydroxylase activity associated with 

fibrogenesis have been shown to increase in rodents with O3 exposure of $0.45 ppm. Some 

studies have shown that this increase persists after exposure stops and that there is an influence 

of exposure pattern on the response. The increased collagen has been correlated with structural 

changes in the lung. Rats exposed to an urban pattern of O3 with daily peaks of 0.25 ppm for 

38 weeks displayed extracellular matrix thickening. Increased levels of collagen in CAR were 

demonstrated in female rats exposed to 0.5 to 1.0 ppm O3 for 6 h/day for 20 months and in 

monkeys exposed to 0.61 ppm for 1 year. Both increased age and health status (e.g., 

emphysemic) were implicated in the increased collagen formation in response to O3 exposure. 

A time-course study (van Bree et al., 2001; Table AX5-5) evaluating the lung injury and 

changes in collagen content in rats exposed acutely or subchronically to 0.4 ppm O3 

demonstrated CAR thickening of septa, which progressed from 7 through 56 days of exposure. 

5-11
Though collagen content decreased with PE recovery, the structural fibrotic changes in ductular 

septa and respiratory bronchioles persisted, suggesting that subchronic O3 exposures in rats 

creates a progression of structural lung injury that can evolve to a more chronic form, which 

includes fibrosis. The biological relevance and adverse health effects of altered protein synthesis 

and collagen accumulation are uncertain. 

5.2.1.6 Differential Gene Expression 

Gohil et al. (2003) examined differential gene expression in C57BL/6 mice exposed to 

1 ppm O3 for three consecutive nights for 8 h (see Table AX5-6). Ozone exposure induced 

changes in expression of 260 genes (80% repressed and 20% induced). Differentially expressed 

genes included those involved in progression of the cell cycle, such as S-adenosyl methionine 

decarboxylase 3 (SAMD), ribonucleotide reductase (RR), and clusterin. Increased transcription 

of these genes suggests O3-induced activation of the cell cycle, with subsequent cellular 

proliferation. This is in accord with the finding of increased epithelial proliferation with acute 

O3 exposure, as discussed in studies in Sections 5.2.4.1 and 5.2.5.1. Several nuclear factor-6B 

(NF-6B)-induced genes were upregulated, including serum amyloid protein, topoisomerase II", 

monocyte chemoattractant protein and platelet-derived growth factor, an inhibitor of apoptosis. 

Upregulation of these genes suggests to the authors that they may account for O3-induced 

proliferation of nonciliated cells and Clara cells. Downregulation of transcripts for isoforms of 

myosins and actins were also observed, which may explain, in part, a mechanism of O3-induced 

vascular permeability. Several members of the CYP family were downregulated, including 2a4, 

2e1, and 2f2, as were aryl-hydrocarbon receptor and several glutathione transferases. 

Metallothionein 1 and 2 and lactotransferrin were upregulated, indicative of their function as 

antioxidants and anti-inflammatory agents. Ozone-induced suppression of immune function is 

suggested by downregulation of transcripts encoding major histocompatibility complex genes, 

lymphocyte-specific proteins, and immunoglobulins. Section 5.2.2.3 discusses the effects of O3 

exposure on the immune system. 

Quinlan et al. (1994) have reviewed the regulation of antioxidant enzymes in lung after 

oxidant injury. A comparison of alterations in gene expression in rat following O3 or hyperoxia 

exposure, both of which induce reactive oxygen species (ROS) and injury to vascular endothelial 

cells and cells of the alveoli, show that both ~1 ppm O3 and 85 to 95% O2 increase expression of 

5-12
CuZnSOD, GSHPx, and catalase. Johnston et al. (1998) also found changes in gene expression 

indicative of inflammation and epithelial injury that occur with hyperoxia in mice (95% O2) that 

are analgous to similar injury that occurs following O3 exposure. 

5.2.1.7 Summary and Conclusions—Biochemical Effects 

Ozone has been shown to interact with a wide range of different cellular components 

including PUFAs, amino acid residues, and some low-molecular-weight compounds (GSH, 

urate, vitamins C and E). As O3 does not penetrate much beyond the ELF, damage likely results 

from its PUFA ozonation products (mostly hydroxyhydroperoxides) involvement in signaling of 

cellular responses such as inflammation. New work has shown that ozonation of PUFA also 

forms the aldehydes nonanal, heptanal, and hexanal, the production of which are dependent on 

AA availability. Saturated phospholipids are thought to reduce the local dose and limit sitespecific cell injury from O3 exposure. Another ozonation product, HNE, creates protein adducts 

that have been linked to apoptosis and heat shock proteins in vitro. 

Both short- and long-term exposures to O3 have been shown to enhance lung xenobiotic 

metabolism, possibly as a result of changes in the number and function of bronchiolar epithelial 

Clara cells and alveolar epithelial Type II cells. This modulation is both species- and regionspecific and includes the isoforms CYP 2B1, CYP 2E1. CCSP is also involved in inflammatory 

responses to O3 exposure. Mouse strains with differing sensitivities to O3 show that responses in 

protein, lactate dehydrogenase (LDH), and inflammatory cell influx are due to CCSP levels and 

changes in lung epithelial permeability. 

Reactions of O3 with AA, GSH, and UA (all antioxidants present in ELF) are a protective 

mechanism. But even with environmentally relevant exposures, the reactivity of O3 is not 

quantitatively quenched, and cell injury occurs in both the lower and upper respiratory tract. 

Early work has shown that short-term exposures to <1 ppm O3 increase antioxidant metabolism. 

Reexposure after a recovery period causes increases equivalent to first-time exposures, 

suggesting that previous exposure is not protective. Elevations in enzyme activity appear to 

increase as a function of age, suggesting that O3 exposure can cause greater lung injury in the 

older animal. Long-term urban patterns of exposure to O3 (daily peaks of 0.25 ppm) caused 

increases in GSHPx and GR, but not SOD. Recent work has suggested that endogenous 

antioxidants moderate the effects of O3 and that this exposure is a subthreshold stimulus for 

5-13
producing effects on peripheral Raw and reactivity, thus indicating a dissociation between 

airway function and inflammation. 

In both acute and short-term studies, a variety of lung lipid changes occur with O3 

exposure, including an increase in AA. With longer exposures (e.g., 0.12 ppm for 90 days), 

an increase in PUFAs and a decrease in cholesterol-esters are seen, indicative of long-term 

alterations of surfactant lipid composition. Whether these changes in lipid content and/or 

metabolism lead to significant changes in surface tension or compliance properties of the lung 

remains unknown. New studies evaluating O3-induced alterations in lipid metabolism have not 

been completed. 

Collagen, a structural protein involved in fibrosis, increases with O3 exposure, and some 

studies have shown that this increase persists after exposure stops. Urban patterns of exposure 

(daily peaks of 0.25 ppm O3 for 38 weeks) created extracellular matrix thickening. Increases in 

centriacinar collagen were demonstrated in female rats exposed to 0.5 to 1.0 ppm O3 for 6 h/day 

for 20 months and in monkeys exposed to 0.61 ppm for 1 year. New work examining the time 

course of lung injury and changes in collagen content in rats exposed acutely or subchronically 

to 0.4 ppm O3 showed centriacinar thickening of septa. Collagen content decreased with PE 

recovery but not the structural fibrotic changes in ductular septa and respiratory bronchioles, 

which suggests that subchronic O3 exposures in rats creates a progression of structural lung 

injury that can evolve to a more chronic form, including fibrosis. 

5.2.2 Lung Host Defenses 

Defense mechanisms, including the mucociliary clearance system, AMs, and humoral- and 

cell-mediated immune system, exist in the lung to protect it from infectious and neoplastic 

disease and inhaled particles. Summaries of key new animal studies examining the effects of O3 

on lung host defenses are presented in Table AX5-7 of Annex AX5. Acute human exposures 

to O3 result in similar effects on AMs (see Chapter 6). 

5.2.2.1 Clearance 

Early studies of O3 effects on the mucociliary escalator showed morphological damage to 

ciliated epithelial cells of the tracheobronchial tree at O3 doses of <1 ppm. Functionally, O3 

slowed particle clearance in rats at doses of 0.8 ppm for 4 h and in rabbits at 0.6 ppm for 2 h 

5-14
exposures. Acute exposures at 0.5 ppm O3 in sheep caused increased basal secretion of 

glycoproteins, while longer exposures reduced tracheal glycoprotein secretions, both of which 

can alter the effectiveness of the mucociliary escalator. Early postnatal exposures of sheep to 

1 ppm O3 caused retardation of normal morphologic development of the tracheal epithelium, 

decreased epithelial mucosa density, decreased tracheal mucous velocity, and delayed 

development of carbohydrate composition. Conversely, alveolar clearance in rabbits after 

acute O3 exposure (0.1 ppm, 2 h/day, for 1 to 4 days) is increased. Longer exposures showed no 

effect, and increased O3 (1.2 ppm) slowed clearance. This pattern of altered clearance also 

occurs in rats. A study using rat tracheal explants exposed to O3 for 10 min (Churg et al., 1996) 

showed that uptake of titanium dioxide (TiO2) and asbestos was enhanced at 0.01 and 0.1 ppm 

O3, respectively. The authors attribute the increased uptake as a direct effect of O3, suggesting 

mediation by H2O2 or hydroxyl radical. Studies of the clearance of the radiolabled chelate 99mTc 

diethylenetriamine pentaacetic acid (99mTc-DTPA) have shown that clearance is significantly 

increased following a 3 h exposure to 0.8 ppm O3 in Sprague-Dawley (SD) rats (Pearson and 

Bhalla,1997). Examination of regional clearance of 99mTc-DTPA in dogs following a 6 h isolated 

sublobar exposure to 0.4 ppm O3 or air showed that O3 decreased the clearance half-time by 50% 

at 1 day following exposure (Foster and Freed, 1999). Clearance was still elevated at 7 days PE 

but had recovered by 14 days. Thus, a single local exposure to O3 increases transepithelial 

clearance but without any influence on contralateral segments, i.e., only for epithelia directly 

exposed to O3. 

Alveolar clearance is slower than tracheobronchial clearance and involves particle 

movement through interstitial pathways to the lymphatic system or movement of particle-laden 

AMs to the bottom of the mucociliary escalator. Exposures of rabbits to 0.1 ppm accelerated 

clearance, whereas 1.2 ppm O3 slowed clearance. A chronic exposure has been shown to slow 

clearance. New evaluations of O3 effects on alveolar clearance have not been performed. 

5.2.2.2 Alveolar Macrophages 

A primary function of AMs is to clear the lung of infectious and noninfectious particles by 

phagocytosis, detoxification, and removal. Further, AMs secrete cellular mediators that recruit 

and activate inflammatory cells in the lungs (see Figure 5-3). Ozone has been shown to inhibit 

phagocytosis at 0.1 ppm for 2 h in rabbits. This inhibition returns to control levels if exposures 

5-15
are repeated for several days. The production of superoxide anion radicals and the activity of 

AM lysosomal enzymes (both involved in bactericidal activity) are inhibited by 3 h exposures to 

0.4 and 0.25 ppm O3 in rodents and rabbits, respectively. Production of interferon-( (IFN-() 

was decreased in rabbit AM by 1 ppm O3 for 3 h. 

New studies have shown that O3 affects AM chemotaxis, cell adhesion, and surface 

expression of cell adhesion molecules (Bhalla, 1996). AM from SD rats exposed to 0.8 ppm O3 

for 3 h showed greater mobility and greater adhesion than air exposed controls. This increased 

mobility and adhesion were attenuated by CD16b and intercellular adhesion molecule-1 

(ICAM-1) antibodies, suggesting that these adhesion molecules modulate O3-induced 

inflammation. Antibodies to tumor necrosis factor-" (TNF-") and IL-1" also mitigated AM 

adherence, suggesting further that the inflammatory response to O3 is mediated by these 

cytokines (Pearson and Bhalla, 1997). Cohen et al. (1996) showed that 1 ppm O3 for 4 h reduces 

binding of INF-( to AM in WEHI-3 cells and, additionally, reduces phagocytic activity, 

production of reactive oxygen intermediates, and elevation of intracellular Ca2+. 

Cohen et al. (2001, 2002) exposed male Fisher 344 (F344) rats to either 0.1 or 0.3 ppm O3 

for 4 h/day, 5 days/week or either 1 or 3 weeks. In this study, superoxide anion production was 

increased at 1 week. Hydrogen peroxide (H2O2) production was reduced at both exposure 

concentrations and durations and was further reduced with INF-( stimulation, suggesting that 

one effect of O3 is compromised killing of bacteria by AM due to the reduction in H2O2 

production. 

Ozone treatment (2 ppm O3, 3 h in female SD rats) caused a time-dependent increase in 

nitric oxide (NO) levels in both AM and type II epithelial cells that was correlated with 

increased expression of inducible nitric oxide synthase (iNOS) mRNA and protein (Laskin et al., 

1998). Inhibition of NF-6B, caused a dose-dependent inhibition of NO and iNOS production. 

Additionally, O3 caused a time-dependent increase in NF-6B binding activity in the nucleus of 

both cell types. The authors hypothesize that O3 exposure causes the cytokines TNF-" and 

IL-1$" to bind to surface receptors and initiate intracellular signaling pathways in AM, leading 

to activation of NF-6B, its entry into the nucleus, and its binding to the regulatory sequences of 

genes such as iNOS to allow their transcription. Additional studies (Laskin et al., 2002), using 

AM isolated from C57Bl6x129 mice with a targeted disruption of the gene for iNOS, showed no 

toxicity to 0.8 ppm O3 for 3 h, as measured by BALF protein levels and nitrotyrosine staining of 

5-16
the lung. Additionally, mice overexpressing human CuZnSOD and mice with a targeted 

disruption of p50 NF-6B were resistant to O3 toxicity. Wild-type mice exposed to O3 showed an 

increase in expression of STAT-1, a protein that binds to the regulatory region of iNOS. Taken 

together, these results suggest to the authors that a number of proteins, including NF-6B, 

phosphoinoside 3-kinase, and STAT-1, that bind to and regulate expression of iNOS are 

modulated by O3 exposure. The same iNOS knockout mice strain exposed to 0.8 ppm O3 for 3 h 

(Fakhrzadeh et al., 2002) showed no increase in AM superoxide anion and prostaglandin. These 

data provide further evidence that NO and its reactive oxidative product, peroxynitrite, are 

important in O3-induced lung injury. Further discussions of the role of nitric oxide 

synthase/reactive nitrogen and cytokines/chemokines in O3-induced inflammation are provided 

in Section 5.2.3. 

5.2.2.3 Immune System 

Other than by natural protection (e.g., opsonizing antibody, nonspecific phagocytosis by 

AM), the immune system defends the lung by mounting three major waves of response: natural 

killer (NK) cells (nonspecific lymphocytes that kill viruses, bacteria, and tumor cells), followed 

by cytotoxic T lymphocytes (TCTL, lymphocytes that lyse specifically recognized microbial and 

tumor-cell targets), followed by antigen-specific antibodies. These T-cell types are involved 

with other immunologically active cells (e.g., B cells and AM), which in a complex manner, 

interact in immunological defense. To date, only a few of these mechanisms have been 

investigated in the context of their role in O3 susceptibility. The effects of O3 on the immune 

system are complex and depend on the exposure parameters and observation periods. T-celldependent functions appear to be more affected than B-cell-dependent functions. Generally, 

there is an early immunosuppressive effect that can, with continued exposure, either return to 

normal or actually enhance immunity. Changes in immune cell population occur with O3 

exposure, including T:B cell ratios in the mediastinal lymph node. Natural killer cell activity 

increases with 1 week exposures of 0.2 to 0.4 ppm O3 but decreases with exposures to 0.82 ppm. 

Ozone exposure has also been shown to be responsible for enhancement of allergic sensitization 

at levels of 0.5 to 0.8 ppm for 3 days. Studies of the effects of O3 on the immune system are 

summarized in Table AX5-7. 

5-17
Garssen et al. (1997) have studied the effects of O3 on non-IgE-mediated pulmonary 

hyper-immune reactions induced by picryl chloride (PCI). BALB/c mice sensitized with PCI, 

both actively and passively (by adoptive transfer of lymphoid cells from pre-sensitized mice), 

were then challenged with picryl sulfonic acid (PSA). The mice were exposed for 12 h to 0.2, 

0.4, or 0.8 ppm O3 during one night, at 4 days or 7 days after skin sensitization done either just 

before or just after PSA challenge (i.e., during the induction or effector phase). Nonsensitized 

mice showed no changes in tracheal reactivity to carbacol with O3 exposure. Sensitized mice 

were hyperreactive to carbachol 48 h after PSA challenge, whereas sensitized mice exposed to 

all concentrations of O3 showed no significant tracheal hyperreactivity to carbachol. The 

sensitized mice also showed a suppressed inflammatory reaction (polymorphonuclear leukocyte, 

PMN) with the 0.8 ppm O3 exposure. Ozone exposure following PSA challenge also caused a 

suppression of tracheal hyperresponsiveness. In a separate experiment wherein mice were 

exposed to O3 before sensitization and then lymphoid cells from these mice were injected into 

nonexposed mice, the recipients also manifested an inhibition of the induction of hyperreactivity. 

These results are opposite to the effect on type I (IgE-mediated) allergic reactions, which the 

authors suggest is due to activation of T-derived lymphocyte helper 2 (Th2) cell-dependent 

reactions that are possibly potentiated by O3 or to a direct effect by O3 on Th1 cells or other cells 

crucial for the tracheal hyperreactivity and inflammation seen in this mouse model. 

Kleeberger et al. (2000, 2001a) have demonstrated a potential interaction between the 

innate and acquired immune system with O3 exposure. Using O3-susceptible (C57BL/6J) 

and O3-resistant (C3H/HeJ) mice, they identified a candidate gene on chromosome 4, Toll-like 

receptor 4 (Tlr4). Ozone exposure (0.3 ppm for 24 to 72 h) of C3H/HeJ and C3H/HeOuJ mice, 

the latter differing from the O3-resistant strain by a polymorphism in the coding region of Tlr4, 

demonstrated greater protein concentrations in the OuJ strain. The two strains exhibited 

differential expression of Tlr4 mRNA with O3 exposure. Thus, a quantitative trait locus on 

chromosome 4 appears to be responsible for a significant portion of the genetic variance in 

O3-induced lung hyperpermeability. In these mouse strains, lavageable protein concentration 

was lowered by inhibition of iNOS and by targeted disruption of Nos2. Comparisons of 

C3H/HeJ and C3H/HeOuJ O3 exposures demonstrated reduced Nos2 and Tlr4 mRNA levels in 

the O3-resistant C3H/HeJ mice. These data are consistent with the hypothesis that O3-induced 

lung hyperpermeability is mediated by iNOS. These studies suggest a role for the Toll-like 

5-18
receptor 4 (TLR4) in the host response to O3 similar to the role it has demonstrated in 

lipopolysaccharide (LPS) sensitivity (Schwartz 2002; Wells et al. 2003). TLR4 signaling is 

thought to be critical to linking the innate and acquired immune system through antigen 

presenting cells and Th1/Th2 differentiation. 

Ozone exposure has been shown to affect antibody responses both in vitro and in mice. 

Becker et al. (1991) demonstrated changes in IgG production in cultured human lymphocytes 

with O3 exposures of 1.0, 0.5, and 0.1 ppm for 2 h. Subsequent to O3 exposure, cells were 

stimulated with pokeweed mitogen (PWM, a T-cell-dependent stimulus) or Staphylococcus 

aureus Cowan 1 strain (SAC, a T-cell-independent stimulus). Both B and T cells were affected 

by O3. T cells also demonstrated an increase in IL-6 and a decrease in IL-2, which suggested to 

the authors that O3 may have direct effects on IgG producing cells and concurrently an effect that 

is mediated by altered production of T cell immunoregulatory molecules. Responses to 

repeated O3 (0.08 to 0.25 ppm) and ovalbumin (OVA, 1%) exposures were compared in “IgEhigh responder” (BALB/c) and “IgE-low responder” (C57BL/6) mice (Neuhaus-Steinmetz et al., 

2000). Ozone appeared to shift the immune response toward a Th2-like pattern in the two mouse 

strains, with differing potentials for developing allergic reactions. 

Another study (Depuydt et al., 2002) demonstrated that O3 (0.1 ppm for 2 h) increases 

allergen-induced airway inflammation in previously sensitized mice but has no effect on the 

sensitization process itself. This study used OVA-pulsed dendritic cells instead of systemic 

adjuvant, which the authors consider a more relevant model of sensitization, as it clearly 

separates the immune response from the challenge and does not obscure regulatory processes as 

does intraperitoneal (i.p.) injections of OVA. They further suggest that dendritic cells, the 

principal antigen-presenting cells in the airway, are an important component of O3-induced 

eosinophilic airway inflammation. 

Surfactant proteins A and D were shown to create an inflammatory feedback loop with 

perturbations in lung immune defenses (reviewed in Hawgood and Poulain, 2001). Earlier 

studies suggested that SP-A is a target for O3 toxicity by causing inhibition of SP-A selfassociation and SP-A-mediated lipid vesicle aggregation. Further, O3 reduced the ability of 

SP-A to inhibit phospholipid secretion by alveolar type II cells and reduced the capacity of SP-A 

to induce superoxide anion production and enhance phagocytosis of herpes simplex virus. 

Bridges et al. (2000) reported that both SP-A and SP-D directly protect surfactant phospholipids 

5-19
and macrophages from oxidative damage by blocking accumulation of thiobarbituric acid 

reactive substances (TBARS) and conjugated dienes. 

Eight human variants of SP-A in Chinese hamster ovary (CHO) cells exposed to O3 (1 ppm 

for 4 h) showed decreased ability to stimulate cytokine (TNF-" and IL-8) production in THP-1 

cells, a macrophage-like cell line (Wang et al., 2002). Each variant had a unique time- and 

dose-dependent pattern of stimulation of cytokine production with O3 exposure, which the 

authors attribute to possible differences in susceptibility to O3 oxidation. Targeted disruption of 

mouse SP-A and SP-D (Hawgood et al., 2002) caused increases in BAL phospholipid, 

macrophage, and protein through 24 weeks of age. Further, the deficient mice developed patchy 

lung inflammation and air space enlargement consistent with emphysema. Future experiments 

using these null mice will help to establish the role of SP-A and SP-D in pulmonary host defense 

to O3 exposure. 

5.2.2.4 Interactions with Infectious Microorganisms 

Ozone-induced dysfunction of host defense systems results in enhanced susceptibility to 

bacterial lung infections. Acute exposures of 0.08 ppm (3 h) O3 can overcome the ability of 

mice to resist infection (by decreasing lung bactericidal activity) with Streptococcal bacteria, 

resulting in mortality. Changes in antibacterial defenses are dependent on exposure regimens, 

species and strain of test animal, species of bacteria, and age of animal, with young mice being 

more susceptible to the effects of O3. The effect of O3 exposure on antibacterial host defenses 

appears to be concentration- and time-dependent. Early studies using the mouse “infectivity 

model,” consisting of exposure to clean air or O3 followed by exposure to an aerosolized 

microorganism, showed that the difference in mortality between O3-exposed groups and controls 

is concentration-related. Chronic exposures (weeks, months) of 0.1 ppm do not cause greater 

effects on infectivity than short exposures, due to defense parameters becoming reestablished 

with prolonged exposures. 

More recent studies of O3-induced modulation of cell-mediated immune responses showed 

effects on the onset and persistence of infection. Cohen et al. (2001, 2002) exposed male F344 

rats subchronically to either 0.1 or 0.3 ppm O3 for 4 h/day 15 days/week, for 1 or 3 weeks. 

Subsequent exposure with viable Listeria monocytogenes demonstrated no observed effect on 

cumulative mortality but did show a concentration-related effect on morbidity onset and 

5-20
persistence. These data suggest that O3 may cause a possible imbalance between Th1 and Th2 

cells, which can subsequently lead to suppression of the resistance to intracellular pathogens. 

Effects of O3 on viral infections are dependent on the temporal relationship between O3 

exposure and viral infection. Only high concentrations (1.0 ppm O3, 3 h/day, 5 days, mice) 

increased viral-induced mortality. No detrimental effects were seen with a 120-day exposure to 

0.5 ppm O3 on acute lung injury from influenza virus administered immediately before O3 

exposure started. But there were O3-enhanced postifluenzal alveolitis and lung parenchymal 

changes. As O3 does not affect lung influenza viral titers, it apparently does not impact antiviral 

clearance mechanisms. In general, the evidence suggests that O3 can enhance both bacterial and 

viral lung infections, but the key mechanisms have not yet been identified. New studies on the 

interactions of O3 and viral infections have not been published. 

5.2.2.5 Summary and Conclusions—Lung Host Defenses 

New data on lung host defenses support earlier work which suggests that mucociliary 

clearance is affected in most test species at just under 1 ppm, with lower levels (~0.1 ppm) 

increasing clearance and somewhat higher levels decreasing clearance. These data also tend to 

suggest mechanisms whereby O3 affects clearance, which include uptake being a direct effect of 

O3, but modulated by ROS and hydroxyl radicals. 

Alveolar macrophage function is disrupted by O3, as shown by several studies showing 

inhibition of phagocytosis at concentrations ranging from 0.1 to 1.2 ppm. This inhibition returns 

to control levels if O3 exposures are repeated for several days. Two new studies corroborate 

earlier findings of increases in AM number in that same exposure range. In this environmentally 

relevant exposure range, new studies support older findings of decreased resistance to microbial 

pathogens, as shown by those endpoints examining superoxide radical formation, altered 

chemotaxis/motility, decreased INF-( levels, decreased lysosomal activity, increased 

prostaglandin E (PGE) levels, and increased NO mRNA and protein. 

New research evaluating the effects of O3 on immune function advances previous work that 

has shown that exposures can enhance or suppress immune responsiveness, depending on the 

species studied, concentration of O3, route of exposure of allergen, and timing of exposure. 

Continuous exposure to O3 impairs immune responses for the first several days of exposure, 

followed by an adaptation to O3 that allows a return of normal immune responses. Most species 

5-21
show little effect of O3 exposures prior to immunization, but show a suppression of responses to 

antigen in O3 exposures post-immunization. Evaluation of mouse strains with geneticallydetermined differential sensitivity or resistance to O3 indicated a possible interaction between the 

innate and acquired immune systems and, further, that O3 may shift the immune response toward 

a Th2-like pattern. Work has also focused on the deleterious effects of O3 exposure on SP-A and 

SP-D and their immunomodulatory function in protecting against oxidative stress. 

Findings from several new studies evaluating the effects of O3 exposures on infectious 

microorganisms are in concurrence with those from previous studies which showed, in general, 

increased mortality and morbidity, decreased clearance, increased bacterial growth, and 

increased severity of infection at exposure levels of 0.1 to 1 ppm O3 for 1 week. 

5.2.3 Inflammation and Lung Permeability Changes 

The normal lung has an effective barrier function that controls bidirectional flow of fluids 

and cells between the air and blood compartments. Ozone disrupts this function, resulting in two 

well-characterized effects of O3 exposure, lung inflammation and increased permeability, which 

are distinct events controlled by independent mechanisms. Ozone initiates inflammation of lung 

tissue by reactions with antioxidants and lipids in ELF (discussed in Section 5.2.1, see 

Figure 5-2). Secondary reaction products generated in this process then cause changes in cell 

membranes, disruption of the lung barrier leading to leakage of serum proteins, influx of PMNs, 

release of bioactive mediators, and movement of compounds from the air spaces into the blood. 

This increased permeability allows accumulation of co-occurring pollutants into the lung tissue. 

The framework for presenting this stereotypical response to O3 consists of discussions covering: 

(1) the time course of these changes, (2) concentration × time (C × T) relationships, 

(3) susceptibility factors, (4) mediators of inflammation, and (5) NO and reactive nitrogen. 

Rats appear to be more resistant to O3-induced inflammation than humans (see Chapter 4). 

With comparable exposure protocols, both species have similar observed inflammatory and 

permeability changes, i.e., controlled human exposure studies discussed in Chapter 6 indicate 

that the majority of acute responses in humans are similar to those observed in animals. 

Ozone also increases the permeability from the air to the blood compartment. Ozone 

(0.8 ppm for 2 h) caused a 2-fold increase of the transport of labeled DTPA from the rat tracheal 

lumen to the blood. This coincided with a 2-fold increase in the number of endocytic vesicles in 

5-22
epithelial cells that contained intraluminally instilled horseradish peroxidase (HRP) as a tracer. 

These studies also suggest an uneven disruption of tight junctions and alternative transport via 

endocytotic mechanisms. In studies aimed at detecting the effects of O3 exposure on regional 

permeability, O3 increased the transmucosal transport of diethylenetriamine pentaacetic acid 

(DTPA) and bovine serum albumin (BSA) more in the trachea and bronchoalveolar zone than in 

the nose. These changes in barrier integrity may allow increased entry of antigens and other 

bioactive compounds (e.g., bronchoconstrictors) into lung tissue. Data from analyses at regular 

intervals PE indicate that maximal increases in BALF protein, albumin and number of PMNs 

occur 8 to 18 h (depending on the study) after an acute exposure to 0.5 to 1.0 ppm O3 ceases. 

Increases in permeability and inflammation have been observed at levels as low as 

0.1 ppm O3 for 2 h/day for 6 days in rabbit and 0.12 ppm in mice (24 h exposure) and rats 

(6 h exposure). After acute exposures, the influence of the time of exposure increases as the 

concentration of O3 increases. The exact role of inflammation in the etiology of lung disease is 

not known; nor is the relationship between inflammation and changes in lung function. 

Table AX5-8 in Annex AX5 summarizes new key studies describing the potential for O3 

exposure effects on lung permeability and inflammation. 

5.2.3.1 Time Course of Inflammation and Lung Permeability Changes 

A study of OVA-sensitized male Dunkin-Hartley guinea pigs exposed to 1.0 ppm O3 for 

3 h showed that levels of PMN significantly increased at 3 h PE, but BAL protein levels did not, 

suggesting a lack of correlation between the two endpoints (Sun et al., 1997). Increased PMN 

without a concordant increase in BAL protein levels were found when the guinea pigs were 

exposed to 1.0 ppm O3 for 1 h and evaluated 24 h PE. The first group also had an increase in 

AHR, but not the second group, which suggests a dissociation between PMN levels and AHR. 

Earlier work demonstrated that O3 exposures of 0.8 to 1 ppm in rat and guinea pig 

transiently increase the permeability from the air to the blood compartment. This permeability is 

greatest in tracheal and bronchoalveolar zones, and may allow increased entry of antigens and 

other bioactive compounds (e.g., bronchoconstrictors) into lung tissues. The time course of the 

influx of PMNs into the lung and the BALF fluid levels of macrophage inflammatory protein-2 

(MIP-2) were found to be roughly similar to that for proteins (Bhalla and Gupta, 2000). 

Adherence of neutrophils to pulmonary vascular endothelium is maximal within 2 h after 

5-23
exposure and returns to control levels by 12 h PE (Lavnikova et al., 1998). Cheek et al. (1995) 

cultured monolayers of rat alveolar type II cells and exposed them to 0.1 or 0.5 ppm O3 for 0.5 h 

to evaluate the effects of O3 on permeability. Permeability increased dose-dependently and the 

higher exposures elicited greater numbers of injured epithelial cells. Exposure to 0.1 ppm O3 

was thought to expedite the restoration of epithelial barrier functions, while at higher exposures, 

neutrophils exacerbated the O3-induced injury. Vesely et al. (1999a) have demonstrated that 

neutrophils contribute to the repair process in O3-injured airway epithelium and they may play a 

role in removal of O3-injured cells. 

Exposures of 3 to 7 days have been found to cause increases in BALF protein and PMNs 

that typically peak after a few days (depending upon species tested and exposures) and return 

toward control even with continuing exposure. Van Bree et al. (2002) observed lower 

BALF levels of protein, fibronectin, IL-6, and inflammatory cells in rats exposed for 5 days to 

0.4 ppm O3 than in rats exposed for 1 day, suggesting adaptation to O3 exposure. Postexposure 

challenge with single O3 exposures at different time points showed recovery of susceptibility 

to O3. McKinney et al. (1998) observed differences in IL-6 levels due to repetitive exposures 

and demonstrated a role of IL-6 in the adaptive response induced by repeated O3 exposures of 

0.5 ppm for 4 h. 

5.2.3.2 Concentration and Time of Exposure 

The relative influence of concentration and duration of exposure (i.e., C × T) has been 

investigated extensively in rats, using BALF protein as an endpoint. Earlier work utilizing 

concentrations of 0.1 to 2 ppm O3 and durations of 1 to 8 h has shown that the interaction 

between C and T is complex. At these levels of exposure, concentration generally dominates the 

response. The smallest C × T product causing an increase in protein was 0.52 ppm • h (0.13 ppm 

× 4 h). C × T studies using the endpoints of changes in lung protein or cell type showed that 

acute damage is a function of cumulative dose. The impact of T is C-dependent (at higher Cs, 

the impact of T is greater); at the lowest C and T values, this dependence appears to be lost. The 

controlled human exposure data described in Chapter 6 concur with most animal data, showing 

that (a) concentration of O3 is an important factor determining O3 responses and (b) duration of 

exposure and ventilation rate also exert substantial influences. 

5-24
New studies evaluating C × T relationships in animal models have not been completed. 

However, a full understanding of C × T relationships in ambient exposures must include the 

recognition that ‘real world’ exposures are cyclic in nature, due to the daily and seasonal 

variations in O3 levels. The concentration of O3, the duration of the exposure, and duration 

between exposures are all relevant to determining the type and level of O3-induced injury. 

5.2.3.3 Susceptibility Factors 

Factors that have been studied for potential impact on the effects of O3 exposure include 

age, gender, nutritional status, exposure to copollutants, exercise, and genetic variability. A full 

characterization of the effects of age on O3 responses has not been completed. Data available 

indicate that effects of age on O3 responses are endpoint-dependent, with young mice, rats, and 

rabbits having greater prostaglandin levels with exposure and senescent rats having greater IL-6 

and N-acetyl-$-D-glucosaminidase levels with exposure. 

A study (Johnston et al., 2000a) compared gene expression of chemokines and cytokine in 

newborn and 8-week-old C57BL/6J mice exposed to 1.0 or 2.5 ppm for 4, 20, or 24 h. The 

newborn mice displayed increased levels of MT mRNA only, while the 8-week-old mice had 

increases in MIP-1", MIP-2, IL-6, and MT mRNA. Comparisons were made with mice of the 

same age groups with exposures to endotoxin (10 ng/mouse for 10 min). Both age groups 

displayed similar cytokine/chemokine profiles with endotoxin exposure. This suggested to the 

authors that the responses to endotoxin (which does not cause epithelial injury) and the 

responses to O3 (which does) demonstrate that differences in inflammatory control between 

newborn and adult mice is secondary to epithelial injury. 

Pregnancy and lactation increased the susceptibility of rats to acute O3, but no clear effects 

of gender have been identified. The effects of vitamin C deficiency on O3 responses are unclear. 

Ascorbate-deficient guinea pigs exposed to O3 demonstrated only minimal effects on injury and 

inflammation (Kodavanti et al., 1995). Utilizing a diet-restricted (20% of the freely fed diet) rat 

model, Elsayed (2001) demonstrated higher survivability upon exposure to higher O3 (0.8 ppm 

continuously for 3 days) compared to freely fed rats. Preexposure to sidestream cigarette 

smoke had been found to cause increased lung injury (Yu et al., 2002). In vitro studies on the 

macrophages from smoke + O3- exposed animals responded by a greater release of TNF-"

5-25
following LPS stimulation when compared to macrophages exposed to air, smoke, or O3 

(0.5 ppm for 24 h) alone. 

Lines of evidence illustrate that genetic background is an extremely important determinant 

of susceptibility to O3. Earlier studies using inflammation-prone (susceptible) C57BL/6J (B6) 

and inflammation-resistant C3H/HeJ (C3) mouse strains and high doses of O3 (2 ppm for 3 h) 

identified Inf-2 as a locus controlling susceptibility. Further studies in these two strains of mice 

using more relevant exposures (0.3 ppm for 72 h) identified that the acute and subacute 

exposures are controlled by two distinct genes, referred to as Inf-1 and Inf-2, respectively 

(Tankersley and Kleeberger, 1994). Exposures to 0.3 ppm O3 for 48 or 72 h, when repeated 

14 days after the initial exposures, caused a smaller increase in BALF protein and numbers of 

macrophages, lymphocytes, and epithelial cells in both strains, but PMN number was greater in 

both strains compared to initial exposure (Paquette et al., 1994). Kleeberger et al. (1997) also 

identified another potential susceptibility gene, tumor necrosis factor (Tnf, which codes for the 

pro-inflammatory cytokine TNF-" ) on a qualitative trait locus on mouse chromosome 17. 

By neutralizing the function of TNF-" with a specific antibody, they were able to confer 

protection against O3 (0.3 ppm, 48 h) injury in susceptible mice. They then demonstrated a role 

for TNF receptor 1 and 2 (TNFR1 and TNFR2, respectively) signaling in subacute (0.3 ppm for 

48 h) O3-induced pulmonary epithelial injury and inflammation (Cho et al., 2001). TNFR1 and 

TNFR2 knockouts were less sensitive to subacute O3 exposure than were WT C57BL/6J mice. 

An integrated and more comprehensive effort to identify the genetic basis for the 

susceptibility to O3-induced lung injury was reported by Savov et al. (2004). In this report, acute 

lung injury in response to high O3 dose (2 ppm for 3 h) was assessed and integrated with 

physiological, biochemical, and genetic observations using nine inbred mouse strains. This work 

indicated the presence of genetic loci on chromosomes 1, 7, and 15 associated with phenotypic 

characteristics for resistance to acute O3-induced lung injury. They identified C3H/HeJ and 

A/J as consistently O3-resistant; C57BL/6J and 129/SvIm as consistently O3-vulnerable; and 

CAST/Ei, BTBR, DBA/2J, FVB/NJ, and BALB/cJ as intermediate in response to O3. 

Ozone-induced changes in CCSP (called CC16 by this group) expression were evaluated in 

five inbred mouse strains: C57BL/6J and CBA both considered sensitive to acute O3-induced 

inflammation, C3H/HeJ and AKR/J both considered resistant, and SJL/J considered intermediate 

(Broeckaert et al., 2003). Two exposure protocols, 1.8 ppm O3 for 3 h or 0.11 ppm O3, 24 h/day 

5-26
for up to 3 days, were used; and BALF and serum were assayed immediately after exposure or at 

6 h PE. Both exposure levels caused a transient increase in CC16 in serum that correlated with 

BALF changes in protein, LDH, and inflammatory cells. There was an inverse relationship 

between preexposure levels of CC16 in BALF and epithelial damage, based on serum CC16 

levels and BALF markers of inflammation. There was also an inverse relationship between 

preexposure levels of albumin in BALF and lung epithelium damage. Based on these results, the 

authors concluded that a major determinant of susceptibility to O3 is due to differences in the 

basal permeability of the airway epithelium. As all of the mouse strains had similar levels of 

preexposure CC16 mRNA, they explored the possible role of CC16 isozymes in differences 

among strains. The CC16 monomer, a 7kD protein, exists in two isoforms with differing pI 

values, CC16a (4.9) and CC16b (5.2). To evaluate the role of CC16 isoform profiles in 

permeability differences between C57BL/6J and C3H/HeJ, this group evaluated the CC16 

protein profiles in BALF of the strains before and after O3 exposure following two-dimensional 

protein electrophoresis analysis. C57BL/6J mice had lower levels of CC16a (the more acidic 

form) than C3H/HeJ. But both the strains had similar levels of CC16b. Based on these 

observations, Broeckaert et al. (2003) concluded that greater epithelial permeability observed in 

C57BL/6J mice may be due to differences in the expression of CC16a and, possibly, other 

antioxidant/anti-inflammatory proteins. 

Wattiez et al. (2003) examined BALF protein from C57BL/6J (O3-sensitive) and 

C3H/HeJ (O3-resistant) mice exposed to filtered air, using a two-dimensional polyacrylamide 

gel approach to analyze the protein profiles. C3H/HeJ mice expressed 1.3 times more Clara cell 

protein16 (CC16) than C57BL/6J mice and, further, expressed more of the acidic isoform of 

CC16. Strain-specific differential expression of isoforms of the antioxidant protein 2 (AOP2), 

the isoelectric point 5.7 isoform in C3H/HeJ and isoelectric point 6.0 isoform in C57BL/6J, were 

observed. These data suggest a protective role for CCSP against oxidative damage and, further, 

that genetic susceptibility to oxidant stress may be moderated, in part, by the gene coding for 

CCSP. Taken together, these mouse studies of genetic susceptibility are useful for helping to 

understand underlying mechanisms potentially leading to O3-induced effects. However, at this 

point, corresponding human polymorphisms have not yet been identified as being associated 

with differing human sensitivities to O3. 

5-27
5.2.3.4 Mediators of Inflammatory Response and Injury 

Ozone reacts with lipids in the ELF or epithelial cell membranes, creating ozonation 

products which then stimulate airway epithelial cells, AMs, and PMNs to release a host of 

pro-inflammatory mediators, including cytokines, chemokines, ROS, eicosanoids, and PAF (see 

Figure 5-3). While neutrophils in the lung characterize an inflammatory response to O3, the 

release of chemotactic mediators by inflammatory cells indicates their state of activation and 

their role in continued inflammation and injury. At O3 exposures of $1 ppm in rodents, these 

mediators recruit PMNs and increase expression of MIP-2 mRNA or BALF levels of MIP-2 

(Driscoll et al., 1993; Haddad et al., 1995; Bhalla and Gupta, 2000). The increased mRNA 

expression was associated with an increased neutrophilia in the lung. Harkema et al. (1987) 

reported neutrophilic rhinitis in monkeys exposed to 0.15 ppm O3 for 6 days. Zhao et al. (1998) 

showed that 0.6 ppm O3 exposure for 2 h in mice and rats causes an increase in monocyte 

chemotactic protein-1 (MCP-1). 

Fibronectin, an extracellular matrix glycoprotein, is thought to have a role in lung 

inflammation and inflammatory disorders, and has been shown to be increased with exposure to 

1 ppm O3 for 14 days. Gupta et al. (1998) observed an increase in both fibronectin protein and 

mRNA expression in the lung of rats exposed to 0.8 ppm O3 for 3 h. A mechanistic role of 

fibronectin in O3-induced inflammation and injury was suggested on the basis of comparability 

of temporal changes in BALF protein, fibronectin, and alkaline phosphatase activity with 

exposures of 1 ppm for 3 h (Bhalla et al., 1999). Studies have reported an effect of O3 on other 

cytokines and inflammatory mediators. An increase occurred for cytokine-induced neutrophil 

chemoattractant (CINC) and NF-6B expression in vivo (Koto et al., 1997); for IL-8 in vivo and 

in vitro (Chang et al., 1998); TNF-", fibronectin, IL-1 and CINC release by macrophages 

ex vivo (Pendino et al., 1994; Ishii et al., 1997); and NF-6B and TNF-" (Nichols et al., 2001; see 

Section 6.9.2 of Chapter 6). An increase in lung CINC mRNA occurred within 2 h after the end 

of a 3 h exposure of rats to 1 ppm O3. The CINC mRNA expression was associated with 

neutrophilia at 24 h PE. Exposure of guinea pig AMs recovered in BALF and exposed in vitro 

to 0.4 ppm O3 for 1 h produced a significant increase in IL-6 and TNF-" (Arsalane et al., 1995). 

An exposure of human AMs to an identical O3 concentration increased TNF-", IL-1$, IL-6, and 

IL-8 and their mRNAs. Ozone exposure (0.3 to 2.5 ppm, 1 to 48 h) of mice caused an increase 

in IL-6, MIP-1", MIP-2, eotaxin, and MT abundance (Johnston et al., 1999a). The IL-6 and MT 

5-28
increase was enhanced in mice deficient in CCSP, suggesting a protective role of Clara cells and 

their secretions (Mango et al., 1998). CCSP deficiency, also increased sensitivity of mice to O3, 

as determined by an increase in abundance of MIP-1" and MIP-2 following a 4 h exposure to 

1.0 ppm O3 (Johnston et al., 1999b). 

Mast cells, which are located below the epithelium, release pro-inflammatory mediators 

and have been shown to contribute to O3-induced epithelial damage. Greater increases in 

lavageable macrophages, epithelial cells and PMNs were observed in mast cell-sufficient mice 

than in mast cell-deficient mice exposed to 0.26 ppm O3 for 8 h per day, 5 days per week 

(Kleeberger et al., 2001b). Increases in inflammatory cells were also observed in mast 

cell-deficient mice, however, O3-induced permeability changes were similar between genotypic 

groups exposed to 0.26 ppm. When an RBL-2H3 mast cell line was exposed to 0.1 to 1.0 ppm 

O3 for 1 h, spontaneous release of serotonin and modest generation of prostaglandin D2 (PGD2) 

occurred only under conditions that caused cytotoxicity (Peden and Dailey, 1995). Additionally, 

O3 inhibited IgE- and A23187-induced degranulation. Mast cells recovered from O3-exposed 

peripheral airways of ascaris-sensitive dogs released significantly less histamine and PGD2 

following in vitro challenge with ascaris antigen or calcium ionophore (Spannhake, 1996). 

Ozone (0.4 ppm, 5 weeks) exposure also promoted eosinophil recruitment in the nose and 

airways in response to instillation of OVA or OVA-pulsed dendritic cells and aggravated allergy 

like symptoms in guinea pigs (Iijima et al., 2001). 

The role of PMNs and cellular mediators in lung injury and epithelial permeability has 

been investigated using antibodies and inhibitors of known specificity to block inflammatory cell 

functions and cytokine activity. Treatment of rats with cyclophosphamide prior to O3 exposure 

(0.8 ppm, 48 h) resulted in a decreased recovery of PMNs in the BALF and attenuated 

permeability induced by O3 (Bassett et al., 2001). 

Pretreatment of animals with antiserum against rat neutrophils abrogated PMN 

accumulation in the lung, but did not alter permeability increase produced by O3. Studies 

utilizing antibodies to selected pro- or anti-inflammatory cytokines suggest a role of TNF-", 

IL-10, and IL-1$ in O3-induced changes in permeability, inflammation and cytokine release 

(Ishii et al., 1997; Reinhart et al., 1999; Bhalla et al., 2002) in exposures of -1 ppm for 3 to 6 h. 

An attenuation of O3-induced increases in permeability and inflammation was also observed in 

mice treated either before or after exposure with UK-74505, a PAF receptor antagonist 

5-29
(Longphre et al., 1999). These results were interpreted to indicate that O3-induced epithelial and 

inflammatory changes are mediated in part by activation of PAF receptors. 

Ozone exposure stimulates macrophage motility toward a chemotactic gradient, and 

macrophages isolated from rats exposed to 0.8 ppm O3 for 3 h adhered to epithelial cells 

(ARL-14) in culture to a greater extent than macrophages from air-exposed controls (Bhalla, 

1996). Both macrophage motility and chemotaxis were attenuated by antibodies to cell adhesion 

molecules CD-11b and ICAM-1, suggesting a role for cell adhesion molecules in O3-induced 

cellular interactions. This may also explain the increased tissue localization and reduced 

recovery of macrophages in BALF (Pearson and Bhalla, 1997) following O3 exposure (0.8 ppm, 

3 h). Studies investigating the mechanisms of PMN recruitment in the lung have explored the 

role of cell adhesion molecules that mediate PMN-endothelial interactions. An exposure of 

female rats to O3 (1 ppm, 2 h) had an attenuating effect on CD-18 expression on AMs and 

vascular PMNs, but the expression of CD62L, a member of selection family, on vascular PMNs 

was not affected (Hoffer et al., 1999). In monkeys, O3-induced (0.8 ppm, 8 h) inflammation was 

blocked by treatment with a monoclonal antibody to CD18, suggesting dependence of PMN 

recruitment on this adhesion molecule (Hyde et al., 1999). Treatment of monkeys with CD18 

antibody also reduced tracheal expression of the $6 integrin (Miller et al., 2001), suggesting that 

lung epithelial cell expression of this adhesion molecule is associated with sites of neutrophil 

recruitment. A single 3 h exposure of rats to 1 ppm O3 caused an elevation in concentration of 

ICAM-1, but not CD-18, in the BALF (Bhalla and Gupta, 2000). Takahashi et al. (1995a) found 

an increase in tissue expression of ICAM-1 in mice exposed to 2 ppm O3 for 3 h, noting a 

temporal correlation of inflammatory activity and ICAM-1 expression, which varied in different 

regions of the lung. A comparable pattern of time-related changes in total protein, fibronectin 

and alkaline phosphatase activity in the BALF of rats exposed to 0.8 ppm O3 for 3 h was also 

noted by Bhalla et al. (1999). Together, these studies support the role of extracellular matrix 

protein and cell adhesion molecules in the induction of lung inflammation and injury. 

5.2.3.5 The Role of Nitric Oxide Synthase and Reactive Nitrogen in Inflammation 

Nitric oxide is a messenger molecule involved in many biological processes, including 

inflammation (see Figure 5-3). Cells in the respiratory tract (including mast cells, neutrophils, 

epithelial cells, neurons, and macrophages) produce three different forms of nitric oxide synthase 

5-30
(NOS), the enzyme that catalyzes the formation of NO. NOS-1(neuronal) and NOS-3 

(endothelial) are constitutively expressed, whereas NOS-2 (also referred to as iNOS) is 

inducible, commonly by pro-inflammatory cytokines. Macrophages isolated from O3-exposed 

(0.8 ppm for 3 h) mice produced increased amounts of NO, superoxide anion, and PGE2, 

but production of these mediators by macrophages from NOS knockout mice was not 

elevated (Fakhrzadeh et al., 2002). Additionally, mice deficient in NOS or mice treated 

with NG-monomethyl-L-arginine, an inhibitor of total NOS, were protected from O3-induced 

permeability, inflammation, and injury, suggesting a role for NO in the production of O3 effects 

(Kleeberger et al., 2001a; Fakhrzadeh et al., 2002). These results contrast with a study showing 

that O3 exposure (of 1 ppm for 8 h/night for 3 nights) produced greater injury, as determined by 

measurement of MIP-2, matrix metalloproteinases, total protein, cell content, and tyrosine 

nitration of whole lung protein, in iNOS knockout mice than in wild-type mice (Kenyon et al., 

2002). This group suggests that protein nitration is related to inflammation and is not dependent 

on iNOS-derived NO. They point out the possible experimental differences, such as O3 

concentration, for inconsistency between their results and those of Kleeberger et al. (2001a). 

Rats that were pretreated with ebselen (a potent anti-inflammatory, immunomodulator, and 

NO/peroxynitrite scavenger) and then were exposed to 2 ppm O3 for 4 h had decreased numbers 

of neutrophils, lowered albumin levels, and inhibited nitration of tyrosine residues in BALF 18 h 

PE, though macrophage iNOS expression was not changed (Ishii et al., 2000a). These results 

suggest an iNOS-independent mechanism for O3-induced inflammation. Jang et al. (2002) 

showed dose-dependent increases in nitrate (indicative of in vivo NO generation) with O3 

exposure (0.12, 0.5, 1, or 2 ppm for 3 h). Functional studies of enhanced pause (Penh) 

demonstrated increases with O3 exposure which were also dose-dependent. Western blot 

analysis of lung tissue showed increases in NOS-1 but not in NOS -3 or iNOS isoforms. These 

results suggest that, in mice, NOS-1 may induce airway responsiveness by a neutrophilic airway 

inflammation. The literature regarding the effects of O3 exposure on NOS activity is complex 

and conflicting. Similarly, the issue of protein nitration as it relates to cell injury due to O3 

exposure is somewhat controversial. 

5-31
5.2.3.6 Summary and Conclusions—Inflammation and Permeability Changes 

Figure 5-3 depicts many of the inflammatory and permeability changes that occur with O3 

exposure. The figure also demonstrates links between inflammatory/permeability responses and 

altered lung function (discussed in Section 5.2.5), ciliary motility (Section 5.2.2.1), AHR 

(Section 5.2.5.3), and possible thrombolytic effects (Section 5.3.3). Airway mucosa in the 

normal lung serves as an effective barrier that controls bidirectional flow of fluids and cells 

between the air and blood compartments. Ozone disrupts this function, resulting in a cascade of 

effects that includes an increase in serum proteins, bioactive mediators, and PMNs in the 

interstitium and air spaces of the lung. Damaged epithelial cells release cytokines, which 

function to recruit and activate AMs and PMNs. PMN recruitment into the lung is maximal at 

several hours PE. PMN recruitment is followed by recruitment of blood monocytes that enter 

the lung and enlarge to become AMs. The AMs persist for days to weeks, phagocytizing injured 

cells. Activated PMNs and AMs continue the cascade of effects by further releasing 

inflammatory mediators, which serve to amplify the initial effects of O3. Generally, the 

initiation of inflammation is an important component of the defense process; however, its 

persistence and/or repeated occurrence can result in adverse health effects. Activation of this 

inflammatory cascade takes several hours. Chemical mediators released early in the cascade 

contribute to effects on pulmonary function. Events occurring later in the cascade, by 

which time O3-induced alterations in pulmonary function have attenuated, are related to 

sustained inflammation. Further, mechanistic separation of inflammation, permeability, and 

AHR is suggested by the temporal disparities of their increases. The O3-induced disruption of 

the tight junctions between epithelial cells also increases the permeability between the air and 

blood compartments. This disruption, occurring with exposures of 0.8 ppm for 2 h, is greater in 

the trachea and bronchoalveolar zone than in the nose and allows entry of particles, including 

bioactive compounds, into lung tissue. 

For environmentally-relevant exposures to O3, exposure concentration dominates the 

response. Studies evaluating C × T relationships have not been published recently. Other 

factors that have been studied for potential impact on the effects of O3 include age, gender, 

nutritional status, genetic variability, exercise and exposure to copollutants. The effects of age 

on lung inflammation are not well known. After an acute exposure to 0.8 or 1 ppm, young mice, 

rats, and rabbits had greater changes in prostaglandins in BALF, but there were no 

5-32
age-dependent effects on BALF protein or cell number. Comparisons of male and female 

animals, and ascorbate deficiency did not reveal significant differences in the effects of O3, but 

exercise during exposure increased susceptibility. 

Important new work has revealed that susceptibility to O3 is, in part, genetically 

determined. Mouse strains with differing sensitivities to O3 have identified genes on separate 

loci controlling various aspects of inflammation, providing additional evidence for the 

mechanistic separation of responses to O3. Such research is summarized in Figure 5-4. 

Kleeberger’s group has identified Inf-1, which modulates acute inflammatory responses; Inf-2, 

which modulates responses to subacute exposures; and TNF-" and TNF receptors, which are 

involved in inflammatory responses. Other research groups have identified loci linked to other 

endpoints. This line of research provides a groundwork for understanding the underlying 

mechanisms of O3-induced injury and can shed light on genes responsible for human 

susceptibility to O3. 

Figure 5-4. Mouse chromosomes on which genes or gene loci have been identified that 

modulate responses to ozone. 

5-33
Recent studies have placed a major focus on mediators released from inflammatory cells to 

understand the mechanisms of O3-induced inflammation and injury. Cytokines and chemokines 

have been shown to be released as a result of stimulation or injury of macrophages, epithelial 

cells, and PMNs. Exposure of guinea pig AMs recovered in BALF and exposed in vitro to 

0.4 ppm O3 produced a significant increase in IL-6 and TNF-". An exposure of human AMs 

to an identical O3 concentration increased TNF-", IL-1$, IL-6, and IL-8. The expression 

of MIP-2 mRNA or BALF levels of MIP-2 increased in mice and rats exposed to O3 

concentrations $1 ppm. An increase after O3 exposure has also been reported for other cytokines 

and inflammatory mediators, including CINC and fibronectin. The CINC mRNA expression 

was associated with neutrophilia at 24 h PE. Ozone exposure of mice also caused an increase in 

IL-6, MIP-1" and eotaxin in mice. Further understanding of the role of mediators has 

come from studies utilizing antibodies and inhibitors of known specificity. In these studies, 

treatment of rats with an anti-IL-6-receptor antibody prior to a nighttime exposure to O3 

abolished O3-induced cellular adaptive response following a subsequent exposure. Studies 

utilizing antibodies to selected pro- or anti-inflammatory cytokines suggest a role for TNF-", 

IL-10 and IL-1$ in O3-induced changes in permeability, inflammation, and cytokine release. 

Studies investigating the mechanisms of PMN recruitment in the lung have explored the 

role of cell adhesion molecules that mediate PMN-endothelial cell interactions. An increase in 

tissue expression of ICAM-1 occurred in mice exposed to 0.8 ppm O3. A comparable pattern of 

time-related changes in total protein, fibronectin, and alkaline phosphatase activity in the BALF 

was observed in rats exposed to 1 ppm O3. In monkeys, the O3-induced inflammation was 

blocked by treatment with a monocolonal antibody to CD18, suggesting dependence of PMN 

recruitment on this adhesion molecule. Together, these studies support the role of extracellular 

matrix protein and cell adhesion molecules in lung inflammation and injury. 

Ozone exposure also affects macrophage functions and, consequently, their role in lung 

inflammation. Macrophages isolated from O3-exposed mice produced increased amounts of 

NO, superoxide anion, and PGE2, but production of these mediators by macrophages from 

NOS knockout mice was not elevated. Additionally, mice deficient in NOS or mice treated 

with NG-monomethyl-L-arginine, an inhibitor of total NOS, were protected from O3-induced 

permeability, inflammation, and injury. These findings suggest a role for NO in the production 

of O3 effects. 

5-34
5.2.4 Morphological Effects 

Most mammalian species show generally similar morphological responses to <1 ppm O3 

that differ only by region, cell type, exposure parameters, and length of time between exposure 

and examination. Constant low exposures to O3 create an early bronchoalveolar exudation that 

declines with continued exposure and drops sharply in the PE period. Epithelial hyperplasia also 

starts early, increases in magnitude for several weeks, plateaus with continuing exposure, and 

declines slowly during PE. Interstitial fibrosis has a later onset, continues to increase throughout 

the exposure, and can continue to increase after the exposure ends. A schematic comparison of 

these duration-response profiles is presented in Figure 5-5 (Dungworth, 1989). Nonhuman 

primates respond more than rats at this concentration, due to differences in antioxidants, the 

CAR (predicted to receive the highest dose of O3), the presence of respiratory bronchioles, acinar 

volume, and differences in the nasal cavity’s ability to “scrub” the O3. Ciliated epithelial cells of 

the airway, Type I epithelial cells of the gas-exchange region, and ciliated cells in the nasal 

cavity are the cells most affected by O3. Ozone-damaged ciliated cells are replaced by 

nonciliated cells, which are unable to provide clearance function, and Type I cells are replaced 

by Type II cells, which are thicker and produce more lipids. Inflammation also occurs, 

especially in the CAR, wherein the tissue is thickened as collagen accumulates. At exposures of 

0.25 ppm O3 (8 h/day, 18 months) in monkeys, the distal airway was found to be remodeled, as 

bronchiolar epithelium replaces the cells present in alveolar ducts. In both rodents and monkeys, 

it appears that the natural seasonal patterns of O3 exposure alter morphology more than 

continuous exposures; thus, long-term animal studies with uninterrupted exposures may 

underestimate morphological effects. 

5.2.4.1 Acute and Subchronic Exposure Effects 

Morphological effects of key acute and subchronic exposure studies are summarized in 

Table AX5-9. Harkema et al. (1997a) reviewed toxicological studies of the nasal epithelial 

response to short-term O3. New information regarding the effects of O3 in this region include 

results indicating that the topical anti-inflammatory corticosteriod fluticasone propionate 

prevents inflammation and mucous cell metaplasia in rats after cumulative O3 exposure (0.5 ppm 

O3, 8 h/day, for 3 or 5 days) (Hotchkiss et al., 1998). Exposure to bacterial endotoxin, a 

common ambient air toxicant, can potentiate mucous cell metaplasia in the nasal transitional 

5-35
Figure 5-5. Schematic comparison of the duration-response profiles for epithelial 

hyperplasia, bronchoalveolar exudation, and interstitial fibrosis in the 

centriacinar region of lung exposed to a constant low concentration of ozone. 

Source: Dungworth (1989). 

epithelium of rats caused by a previous 3 day 0.5 ppm O3 exposure (Fanucchi et al., 1998). Male 

F344/N Hsd rats were intranasally instilled with endotoxin after exposure to filtered air (FA) or 

0.5 ppm O3, (8 h/day for 3 days). Mucous cell metaplasia was not found in the air/endotoxin 

group but was found in the O3/saline group and was most severe in the O3/endotoxin group. A 

similar synergistic effect was demonstrated by Wagner et al. (2001a,b) with exposure of Fischer 

rats to O3 and/or endotoxin for 8 h per day for 3 days. Ozone alone created epithelial lesions in 

the nasal transitional epithelium, while endotoxin alone caused lesions in the epithelium of the 

nose and conducting airways. The enhanced O3-induced mucous cell metaplasia was related to 

neutrophilic inflammation. 

Pre-metaplastic responses, such as mucin mRNA upregulation, neutrophilic inflammation, 

and epithelial proliferation, were shown to be responsible for O3-induced mucous cell metaplasia 

in the transitional epithelium of rats (Cho et al., 1999a, 2000). Male F344/N rats exposed to O3 

(0.5 ppm, 8 h/day for 1, 2, or 3 days) showed a rapid increase in an airway-specific mucin gene 

5-36
mRNA after exposure to O3, both before and during the onset of mucous cell metaplasia. 

Neutrophilic inflammation coincided with epithelial DNA synthesis and upregulation of 

rMuc-5AC but was resolved before the development of epithelial metaplasia. The mucous cell 

metaplasia was neutrophil-dependent, whereas O3-induced epithelial cell proliferation and mucin 

gene upregulation were neutrophil-independent. 

Dormans et al. (1999) compared the extent and time course of fibrotic changes in mice, 

rats, and guinea pigs exposed to 0.2 or 0.4 ppm O3 for 3, 7, 28, and 56 days. They found a 

concentration-related centriacinar inflammation in all three species, with a maximum after 

3 days of exposure and total recovery within 3 days after exposure. Repair of O3-induced 

damage by removal of injured epithelial cells is enhanced by the influx of neutrophils (Hyde 

et al., 1999; Veseley et al., 1999b; Miller et al., 2001; see Section 5.2.3). A study examining the 

kinetics of early cellular responses to O3 utilized bromodeoxyuridine to label S-phase cells 

(Hotchkiss et al., 1997). Labeling indices for rat nasal transitional epithelial cell DNA were 

greatest 20 to 24 h after O3 (0.5 ppm for 8 h) exposure, and remained greater than control at 

36 h PE. 

Very few published studies have explicitly explored susceptibility factors such as species, 

gender, age, antioxidant defense, acute and chronic airway disease, and exercise. Most typical 

laboratory species studied have qualitatively similar effects associated with O3 exposure. 

Dormans et al. (1999) compared morphological, histological, and biochemical effects in the rat, 

mouse, and guinea pig following O3 exposure and recovery in clean air. Wistar RIV:Tox male 

rats, specific-pathogen free (SPF)-bred National Institutes of Health (NIH) male mice, and 

Hartley Crl:(HA)BR male guinea pigs were continuously exposed to FA, 0.2 or 0.4 ppm for 3, 7, 

28, and 56 days. Recovery from 28 days of exposure was studied at intervals of 3, 7, and 28 

days PE. The mouse was the most sensitive as shown by a concentration and exposure 

time-dependent persistence of bronchiolar epithelial hypertrophy, elevated lung enzymes, and 

slow recovery from exposure. Exposure to the high dose for 56 days in both rats and guinea pigs 

caused increased amounts of collagen in ductal septa and large lamellar bodies in Type II cells. 

The inflammatory response was greater in the guinea pig. Overall, the authors rated mice as 

most susceptible, followed by guinea pigs and rats. However, this ranking was done without 

first adjusting for differences in species-specific delivered dose of O3. 

5-37
Ferrets, monkeys, and rats were exposed to O3 (1.0 ppm, 8 h) to compare airway effects 

(Sterner-Kock et al., 2000). The ferrets and monkeys had similar epithelial necrosis and 

inflammation that was more severe than that found in rats. Because ferrets have a similar 

pulmonary structure as humans (e.g., well-developed respiratory bronchioles and submucosal 

glands), the authors concluded that the ferret would be a better model than rodents 

for O3-induced airway effects. Age susceptibility is dependent on the endpoint examined 

(see Chapter 4 for discussions of age-related differences in O3 dosimetry). One study (Dormans 

et al., 1996) demonstrated that O3-induced centriacinar lesions are larger in younger rats than in 

older rats with exposures to 0.4 ppm O3 for 1 to 7 days. 

New studies have examined O3-induced morphological effects in compromised laboratory 

animals. Rats with endotoxin-induced rhinitis were more susceptible to mucous cell metaplasia 

in the nasal transitional epithelium caused by a 3 day exposure to 0.5 ppm O3 (Cho et al.,1999b). 

Wagner et al. (2002) reported a similar O3-induced enhancement of inflammatory and epithelial 

responses associated with allergic rhinitis. Brown Norway rats were exposed to 0.5 ppm O3, 

8 h/day for 1 day or 3 consecutive days and then immediately challenged intranasally with either 

saline or OVA. Multiple exposures to O3 caused greater increases in mucosubstances produced 

in the nose by allergen challenge. 

Recent research has focused on the concept of O3-susceptible and -nonsusceptible sites 

within the respiratory tract, including in situ antioxidant status and metabolic activity. Plopper 

et al. (1998) examined whether the variability of acute epithelial injury to short-term O3 

exposure within the tracheobronchial tree is related to local tissue doses of O3 or to local 

concentrations of GSH. Adult male rhesus monkeys exposed to O3 (0.4 or 1.0 ppm for 2 h) 

demonstrated significant cellular injury at all sites, but the most damage, along with increased 

inflammatory cells, occurred in the proximal respiratory bronchiole. A significant increase in 

GSH was found in the proximal bronchus at 0.4 ppm O3 and a decrease in the respiratory 

bronchiole at 1.0 ppm O3. A significant decrease in the percentage of macrophages, along with 

significant increases in the percentage of neutrophils and eosinophils, and a doubling of total 

lavage protein, were found after exposure to 1.0 ppm O3 only. The authors concluded that the 

variability of local O3 dose in the respiratory tract was related to inhaled O3 concentration and 

was closely associated with local GSH depletion and with the degree of epithelial injury. 

5-38
Plopper and colleagues (e.g., Watt et al., 1998; Paige et al., 2000a) explored the sitespecific relationship between epithelial effects of O3 exposure and the metabolism of 

bioactivated compounds within the respiratory tract of rats. The distribution of CYP 2E1­

dependent activity, measured with a selective substrate (p-nitrocatechol), was found to be 

highest in the distal bronchioles and minor daughter airways and lower in the lobar bronchi and 

major daughter airways. Short-term O3 exposure (1 ppm for 8 h) increased CYP 2E1 activity in 

the lobar bronchi/major daughter airways only; however, long-term O3 exposure (1 ppm for 

90 days) decreased CYP 2E1 activity in the major and minor airways, further complicating the 

interpretation of O3 effects based on concentration and duration of exposure and recovery. Rats 

treated i.p. with 1-nitronaphthalene, a pulmonary toxicant requiring metabolic activation, and 

exposed to 0.8 ppm O3, 8 h/day for 90 days showed greater histopathologic and morphometric 

effects in the CAR of the lung (Paige et al., 2000b). Despite reported tolerance to oxidant stress 

after long-term O3 exposure, there was increased severity of ciliated cell toxicity. 

5.2.4.2 Summary of Acute and Subchronic Morphological Effects 

Short-term exposures to O3 cause similar alterations in lung structure in a variety of 

laboratory animal species. Cells in the CAR are the primary targets of O3, but ciliated epithelial 

cells in the nasal cavity and airways and Type I epithelial cells in the gas exchange region are 

also targets. New work has shown that a topical anti-inflammatory corticosteroid can prevent 

these effects in nasal epithelia, while exposure to bacterial endotoxin can potentiate the effects. 

Ozone-induced fibrotic changes in the CAR are maximal at 3 days of exposure and recover 

3 days PE with exposures of 0.2 ppm in rodents. New studies of susceptibility factors 

demonstrated that ferrets and monkeys have similar inflammatory and necrotic responses to 

1 ppm O3, which differs from lesser injury seen in rats. Rats with induced allergic rhinitis are 

more susceptible to 0.5 ppm O3 than are control animals. Important new work has demonstrated 

variability of local O3 dose and subsequent injury in the respiratory tract due to depletion of 

GSH. The proximal respiratory bronchiole receives the most acute epithelial injury from 

exposures #1 ppm, while metabolic effects were greatest in the distal bronchioles and minor 

daughter airways. Table 5-1 summarizes new studies evaluating the morphological effects of a 

single acute O3 exposure. 

5-39
3

Table 5-1. Summary of Studies that Evaluated Morphological Effects 

of a Single Acute O3 Exposure. 

l) 

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0.25 20-90 Rat nr nr 8 88 nr 8 nr nr Postlethwait et al. 

min (2000) 

0.5 " " nr nr 88 8 nr 8 nr nr Postlethwait et al. 

(2000) 

1 " " nr nr 888 8 nr 8 nr nr Postlethwait et al. 

(2000) 

1 90 min Rat nr nr 888 88 8 8 nr nr Joad et al. (2000) 

0.5 8 hr Rat 888 nr nr nr nr nr nr nr Hotchkiss et al. 

(1997) 

0.5 8 hr Rat 888 nr nr nr nr nr nr nr Cho et al. 

(1999a, 2000) 

0.5 8 hr Rat 88 nr nr nr nr nr nr nr Wagner et al. 

(2002) 

1 8 hr Rat nr nr 8 nr nr 888 nr nr Schelegle et al. 

(2001) 

1 8 hr Rat 88 nr 8 8 nr 8 nr nr Vesely et al. 

(1999a) 

1 8 hr Rat nr nr nr nr nr 8 8 nr Sterner-Kock 

et al. (2000) 

0.4 2 hr Monkey nr ÷ 8 8 nr 88 nr nr Plopper et al. 

(1998) 

1 " " nr 8 8 8 nr 88 nr nr Plopper et al. 

(1998) 

0.8 8 hr Monkey nr nr nr nr nr 88 nr nr Hyde et al. (1999) 

1 8 hr Monkey nr nr nr nr nr 88 nr nr Sterner-Kock 

et al. (2000) 

1 8 hr Ferret nr nr nr nr nr 88 nr nr Sterner-Kock 

et al. (2000) 

Symbols: ÷ = no change; 8 = small increase in indices reflecting O3-induced morphological changes in specified upper 

or lower respiratory region; 88 = relatively moderate morphological changes induced by O3 exposure; 888 = marked 

morphological changes or remodeling due to O3 exposure; nr = changes not reported. The number of arrows in each study is 

based on the authors’ description of injury or from graphs in the paper comparing O3-exposed animals to air-exposed controls. 

As endpoints for the studies varied among papers and information about injury was largely descriptive, the table does not 

attempt a quantitative comparison, but only a subjective illustration of the current literature. Specific types of changes are 

discussed in text of Section 5.2.4. 

5-40
5.2.4.3 Subchronic and Chronic Exposure Effects 

Summaries of new studies of morphological effects of subchronic and chronic exposures 

are listed in Table AX5-10 in Annex AX5. In general, as the duration of exposure lengthens, 

there is no concomitant linear increase in the intensity of effect for any given endpoint. Rather, 

as exposure proceeds past 1 week to 1 year, Type I cell necrosis and inflammatory responses 

generally decrease to near control values, and hyperplastic and fibrotic changes remain elevated. 

After long-term exposure ended, some indicies of fibrosis persisted and, in some cases, became 

more severe during PE periods in clean air. 

Effects of O3 on the upper respiratory tract of F344 rats exposed to O3 (0.12, 0.5, or 

1.0 ppm for 20 months) included marked mucous cell metaplasia in the rats exposed to 0.5 and 

1.0 ppm O3, but not at 0.12 ppm O3 (Harkema et al., 1997a). In a follow-up study, hyperplasia 

was found in the nasal epithelium of rats exposed to 0.25 and 0.5 ppm, 8 h/day, 7 days/week, for 

13 weeks (Harkema et al., 1999). The mucous cell metaplasia, and associated intraepithelial 

mucosubstances, induced by 0.5 ppm O3 persisted for 13 weeks after exposure. An acute (8 h) 

exposure to 0.5 ppm O3 13 weeks after the chronic exposure induced an additional increase of 

mucosubstances in the nasal epithelium of rats but not in rats chronically exposed to 0 or 

0.25 ppm O3. The persistent nature of the O3-induced mucous cell metaplasia in rats reported in 

this study suggests that O3 exposure may have the potential to induce similar long-lasting 

alterations in the airways of humans. 

No significant changes in nasal tissue were seen in rats continuously exposed for 49 days 

to the ambient air of Mexico City, Mexico (Moss et al., 2001). A rat study using 6-month 

exposures to ambient air of Sao Paulo, with a different pollutant composition than that of 

Mexico City, demonstrated development of secretory hyperplasia in rats (Lemos et al., 1994). 

However, without information on specific differences in ambient pollution composition in the 

two cities, the studies cannot be compared; nor can the observed changes be confidently 

attributed to O3 versus the overall oxidant-contaminated ambient air mix. Because of the 

persistent nature of these changes in the controlled studies with rats and the fact that the upper 

airways of humans are probably more sensitive (like in the monkey) the authors suggested that 

long-term exposure to ambient levels of O3 could induce significant nasal epithelial lesions that 

may compromise the upper respiratory tract defense mechanisms of exposed human populations. 

5-41
Rats exposed to 0.5 ppm O3 for 1 month exhibited Bcl-2 in protein extracts of nasal 

epithelium (Tesfaigzi et al., 1998). Further, after 3 and 6 months of exposure, the number of 

metaplastic mucous cells in the transitional epithelium was indirectly related to the percentage of 

cells that were Bcl-2 positive. Cells from rats exposed to FA did not express any Bcl-2. This 

study suggests that apoptosis regulators like Bcl-2 may play a role in the development and 

resolution of mucous cell metaplasia in the nasal airway. 

A spectrum of lesions was reported (Herbert et al., 1996) in the nasal cavity and 

centriacinar lung of male and female mice exposed to 0.5 or 1.0 ppm of O3 for 2 years; the 

lesions persisted with continued exposure for 30 months. These lesions included bone loss in the 

maxilloturbinates, mucosal inflammation, mucous cell metaplasia in the nasal transitional 

epithelium and increased interstitial and epithelial thickening in the proximal alveolar region. 

In the CAR, there were increased numbers of nonciliated cells. However, changes in other 

endpoints including lung function and lung biochemistry were not evident. The investigators’ 

interpretation of the entire study is that rodents exposed to the two higher O3 concentrations had 

some of the structural hallmarks of chronic airway disease in humans. 

A chronic study using a simulated, seasonal O3-exposure pattern was reported by Plopper 

and colleagues (Evans et al., 2003; Schelegle et al., 2003a; Chen et al., 2003; Plopper and 

Fanucchi, 2000). Infant rhesus monkeys (30 days old) were exposed to FA, house dust mite 

allergen aerosol (HDMA), or O3 + HDMA. The 0.5 ppm O3 exposures were 8 h/day for 5 days, 

every 14 days for a total of 11 O3 episodes. Half of the monkeys were sensitized to HDMA 

(Dermatophagoides farinae) at 14 and 28 days of age. The sensitized monkeys were exposed to 

HDMA for 2 h/day on Days 3 to 5 of the FA or O3 exposures. The lungs were removed during 

the last FA exposure and the right and left cranial and right middle lobes were separately 

inflation fixed. Microdissection and morphometric analyses were performed on the conducting 

airways to the level of the most proximal respiratory bronchiole. Repeated exposures to O3 or O3 

+ HDMA over a 6-month period resulted in an atypical development of the basement membrane 

zone (BMZ) of airways in nonsensitized developing monkeys. Remodeling in the distal 

conducting airways was found in the sensitized monkeys as a result of the damage and repair 

processes occurring with repeated exposure (Evans et al., 2003; Schelegle et al., 2003a). Lung 

function changes in these monkeys (Schelegle et al., 2003b), and associated adaptation of the 

respiratory motor responses (Chen et al., 2003), are described in Section 5.2.5.2. Collectively, 

5-42
these findings provide a plausible pathophysiologic basis for changes in airway function 

described in children growing up in polluted metropolitan areas (e.g., Tager, 1999) 

(see Chapter 7). 

Necropsy of the left caudal lobe of these infant monkeys showed accumulation of mucous 

cells and eosinophils within the combined epithelial and interstitial compartments in the 

conducting airways and in the terminal/respiratory bronchioles (Schelegle et al., 2003a). House 

dust mite sensitization and HDMA challenge alone, or combined with O3 exposure, resulted in 

significantly greater eosinophil accumulation in the conducting airways when compared to FAand O3-only exposures. A significant accumulation of eosinophils was found in the 

terminal/respiratory bronchioles of the sensitized monkeys challenged with HDMA when 

compared to monkeys exposed to FA, O3, and HDMA + O3. The mean mass of mucous cells 

increased in the fifth generation conducting airways of sensitized animals challenged with 

HDMA alone and when combined with O3 exposure and in the terminal bronchioles of sensitized 

animals exposed to HDMA + O3. The tracheal basement membrane of HDMA-sensitized 

monkeys exposed to HDMA or to HDMA + O3 was significantly increased over controls; 

however, there were no significant changes in the airway diameter of proximal and mid-level 

airways. Exposures of sensitized young monkeys to HDMA alone, or to O3 alone, resulted in 

eosinophilia of the mid-level conducting airways and the terminal/respiratory bronchioles, but 

without alterations in airway structure or function. The authors interpreted these findings to 

indicate that the combination of cyclic O3 exposure and HDMA challenge in HDMA-sensitized 

infant monkeys acts synergistically to produce an allergic-reactive airway phenotype 

characterized by significant eosinophilia of mid-level conducting airways, transmigration of 

eosinophils into the lumen, and an altered structural development of conducting airways that is 

associated with increased Raw and nonspecific airway reactivity (see Section 5.2.5). 

Examination of development of the tracheal BMZ in these monkeys (Evans et al., 2003) 

showed that with exposures to either O3 or HDMA + O3, BMZ development was affected. 

Abnormalities in the BMZ included: (1) irregular and thin collagen throughout the BMZ; 

(2) perlecan depleted or severely reduced; (3) fibroblast growth factor receptor-1 (FGFR-1)

immunoreactivity reduced; (4) fibroblast growth factor-2 (FGF-2) immunoreactivity absent in 

perlecan-deficient BMZ, but present in the lateral intercellular space (LIS), in basal cells, and in 

attenuated fibroblasts; (5) syndecan-4 immunoreactivity increased in basal cells. The authors 

5-43
interpret these data as suggesting that O3 target cells are associated with synthesis of epithelial 

BMZ perlecan. The absence of FGF-2, normally stored in the BMZ, could affect downstream 

signaling in airway epithelium and could be responsible for the abnormal development of the 

airway seen in this study and thus be an important mechanism modulating O3-induced injury. 

Mid-level bronchi and bronchioles from these monkeys (Larson et al., 2004) demonstrated 

decrements in the density of epithelial nerves in the axial path between the sixth and seventh 

airway generations in exposures to O3. Combined O3 + HDMA exposures exacerbated this 

reduction. They attribute this loss of nerve plexuses to neural regression or stunted nerve 

development, the latter corroborated by the Evans et al. (2003) finding of decreased growth 

factors following O3 exposure. Additionally, they found streaks or clusters of cells that were 

immunoreactive for protein gene product 9.5 (PGP 9.5, a pan-neuronal marker) and negative for 

calcitonin gene-related peptide. The functional significance of this is unknown but suggests to 

the authors a possible injury-repair process induced by O3. 

Remodeling of the distal airways and CAR is one of the most disturbing aspects of the 

morphological changes occurring after subchronic and chronic exposure to O3. Recently, 

bronchiolization was reported in rats exposed to 0.4 ppm O3 for only 56 days (van Bree et al., 

2001). They also found collagen formation progressively increased with increasing O3 exposure 

and persisted into PE recovery. In addition to centriacinar remodeling, Pinkerton et al. (1998) 

reported thickening of tracheal, bronchial, and bronchiolar epithelium after 3 or 20 months 

exposure to 1 ppm O3, but not to 0.12 ppm. Although some older literature had reported that 

chronic exposures to #1.0 ppm O3 cause emphysema, none of the more recent literature supports 

this hypothesis. 

5.2.4.4 Summary and Conclusions—Subchronic and Chronic Morphological Effects 

The progression of effects during and after a chronic O3 exposure across a range of 0.5 to 

1.0 ppm is complex, with inflammation peaking over the first few days of exposure, then 

dropping, next plateauing, and then finally largely disappearing. Epithelial hyperplasia follows a 

somewhat similar pattern. Effects of 0.5 ppm O3 for 20 months on the nasal mucosa include 

atrophy of nasal turbinates, epithelial hyperplasia, and mucous cell metaplasia, which persisted 

long after the exposure ceased. Fibrotic changes in lung tissue increase very slowly over months 

of exposure; and, after exposure ceases, the changes sometimes persist or increase. The pattern 

5-44
of exposure in this same concentration range determines effects, with 18 months of daily 

exposure causing less morphologic damage than exposures on alternating months. This is 

important, given that environmental O3 exposure is typically seasonal. Plopper and colleagues’ 

long term study of infant rhesus monkeys exposed to simulated, seasonal O3 (0.5 ppm 8 h/day 

for 5 days, every 14 days for 11 episodes) demonstrated: (1) remodeling in the distal airways; 

(2) abnormalities in tracheal basement membrane; (3) eosinophil accumulation in conducting 

airways; and (4) decrements in airway innervation. These findings advance earlier information 

regarding possible injury-repair processes occurring with seasonal O3 exposures. 

5.2.5 Effects on Pulmonary Function 

5.2.5.1 Acute and Subchronic Exposure Effects on Pulmonary Function 

Numerous pulmonary function studies of the effects of acute O3 exposure (defined here 

as #1 week of exposure) in several animal species have been conducted and generally show 

responses similar to those of humans (e.g., increased breathing frequency, decreased tidal 

volume, increased resistance, decreased forced vital capacity [FVC] and changes in the 

expiratory flow-volume curve). These effects are seen at 0.25 to 0.4 ppm O3 for several hours in 

a number of species. At concentrations of $1 ppm, breathing mechanics (compliance and 

resistance) are affected. The breathing pattern returns to normal after O3 exposure. In rats 

exposed to 0.35 to 1 ppm O3 for 2 h/day for 5 days, there was a pattern of attenuation of 

pulmonary function responses similar to that observed in humans. Concurrently, there was no 

attenuation of biochemical indicators of lung injury or of morphological changes. 

Work demonstrating attenuation of pulmonary function effects (see Table AX5-11) was 

completed by Wiester et al. (1996), who exposed male Fischer 344 rats to 0.5 ppm O3 for either 

6 or 23 h/day over 5 days. Ozone-induced changes in lung volume were attenuated during the 

5 exposure days and returned to control levels after 7 days recovery. The responses to repeated 

O3 exposure in rats were exacerbated by reduced ambient temperature, presumably as a result of 

increased metabolic activity. 

In an attempt to model susceptible populations, researchers have utilized inbred mouse 

strains with varying ventilatory responses to O3. As differences were seen in inflammatory 

responses to acute O3 exposures in C57BL/6J and C3H/HeJ mice, comparisons were also made 

of their ventilatory responses (Tankersley et al., 1993). Following an exposure of 2 ppm O3 for 

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3 h, breathing frequency (f 0 B), tidal volume (VT), and minute ventilation (VE) were measured 1 

and 24 h in both normocapnia (or air at ~0% CO2) and hypercapnia (5 or 8% CO2). They 

demonstrated that acute O3 exposures caused altered hypercapnic ventilatory control, which 

varied between strains. This suggested to the authors that O3-induced alterations in ventilation 

are determined, at least in part, by genetic factors. A caveat regarding studies such as this using 

high exposure concentrations is that events observed at high concentrations may differ from 

those observed at near-ambient O3 levels. 

Paquette et al. (1994) measured ventilatory responses in C57BL/6J and C3H/HeJ mice 

given repeated acute exposures of 0.3 ppm for 48 and 72 h. The two strains had differing 

0 responses to both normocapnia and hypercapnia. Normocapnic VE was greater following 

subacute O3 exposure in C57BL/6J mice than in C3H/HeJ mice, due to increased fB and 

reduced VT, respectively. This suggests that the increased VT in C57BL/6J mice may contribute 

to the increased susceptibility to lung injury due to a greater dose of O3 reaching the lower lung. 

Hypercapnic ventilatory responses following subacute O3 exposures demonstrated reduced VE 

(due to decreased VT) in the C57BL/6J mice only. Evaluations of O3 dosimetry were performed 

in these two strains using 18O3-labeled ozone (2 ppm for 2-3 h) (Slade et al., 1997). Immediately 

after exposures of 2 ppm 18O3 for 2 to 3 h, C3H/HeJ mice had 46% less 18O in lungs and 61% 

less in the trachea than C57BL/6J. Additionally, C3H/HeJ mice had a greater body temperature 

decrease following O3 exposure than C57BL/6J mice, suggesting that the differences in 

susceptibility to O3 are due to differences in the ability to decrease body temperature and, 

consequently, decrease the dose of O3 to the lung. 

Tracheal transepithelial potential has also been shown to differ in eight mouse strains 6 h 

after exposure to 2 ppm O3 for 3 h (Takahashi et al., 1995b). AKR/J, C3H/HeJ, and CBA/J were 

identified as resistant strains; and 129/J, A/J, C57BL/6J, C3HeB/FeJ, and SJL/J were identified 

as susceptible strains. The authors noted that strains’ responses in this parameter did not show 

concordance with inflammatory responses, suggesting to the authors that the two phenotypes are 

not controlled by the same genetic factors. 

Savov et al. (2004) characterized ventilatory responses in nine mouse strains exposed to O3 

(2.0 ppm O3 for 3 h). The C57BL/6J strain was hyporeactive to methacholine (MCh) prior to O3, 

but was very responsive to MCh following O3. Conversely, C3H/HeJ mice had an intermediate 

5-46

0
baseline Penh and a small response to MCh following O3 exposure. This study corroborates the 

evidence of no consistent relationship between respiratory Penh and inflammation. 

5.2.5.2 Summary and Conclusions—Acute and Subchronic Effects on 

Pulmonary Function 

Early work has demonstrated that during acute exposure of ~0.2 ppm O3 in rats, the most 

commonly observed alterations are increased frequency of breathing and decreased tidal volume 

(i.e., rapid, shallow breathing). Exposures of ~1.0 ppm O3 affect breathing mechanics 

(compliance and resistance). Additionally, decreased lung volumes are observed in rats with 

acute exposures at levels of 0.5 ppm. New work utilizing inbred mouse strains with varying 

ventilatory responses to O3 has suggested that: (1) control of the ventilatory response is 

determined, at least in part, by genetic factors; (2) increased VT in some strains may contribute 

to lung injury due to a greater dose of O3 reaching the lower lung; (3) the ability to reduce body 

temperature in some strains may account for their decreased O3-induced lung injury; and 

(4) tracheal transepithelial potential is determined, in part, by genetic factors. Importantly, the 

genetic loci that appear to be modulating various aspects of pulmonary responses to O3 differ 

from each other and from loci controlling inflammatory responses. 

Exposures of 2 h/day for 5 days create a pattern of attenuation of pulmonary function 

decrements in both rats and humans without concurrent attenuation of lung injury and 

morphological changes, indicating that the attenuation did not result in protection against all the 

effects of O3. Chronic O3 exposure studies evaluating pulmonary function are not available. 

Earlier work has demonstrated that repeated daily exposure of rats to an episodic profile of O3 

caused small, but significant, decrements in lung function that were consistent with early 

indicators of focal fibrogenesis in the proximal alveolar region without overt fibrosis. 

5.2.5.3 Ozone Effects on Airway Responsiveness 

Effects of O3 on airway reactivity have been observed in a variety of species across an 

exposure range of 0.5 to 1 ppm. Many of the new studies on pulmonary function in laboratory 

animals allow a better prediction of the effects of O3 exposure on the exacerbation of asthma 

symptoms and the risk of developing asthma in humans. However, it is necessary to understand 

the factors that determine airway responsiveness across different mammalian species, as is 

discussed in Chapter 4. 

5-47
Traditional studies of airway responsiveness require sedation in both infants and laboratory 

animals. Laboratory animal studies employ intravenous agonist challenges as well as inhalation 

challenges, though inhaled agonist challenges are preferred in humans. Exercise testing is 

not possible with sedation unless exercise is “simulated” by increasing ventilation using 

elevated CO2, and the need for artificial ventilation in laboratory animal studies may cause 

breathing patterns that affect O3 deposition. Joad et al. (2000) reported that when 1 ppm O3 for 

90 min is administered to isolated rat lung at either 2.4 mL/40 breaths/minute (bpm) or 

1.2 mL/80 bpm, the more rapid breathing pattern elicits less epithelial cell injury than the slower 

breathing pattern. Though this study design does not really model the rapid, shallow breathing 

elicited in the intact animal, it shows greater reduction in injury in the proximal axial airway 

compared to its adjacent airway branch and terminal bronchiole. The rapid, shallow breathing 

pattern protects the large conducting airways of rats but causes a more even distribution of 

epithelial cell injury to the terminal bronchioles (Schelegle et al., 2001). Postlethwait et al. 

(2000) demonstrated that the conducting airways are the primary site of acute cytotoxicity from 

O3 exposure. Three-dimensional mapping of the airway tree in SD rat isolated lung exposed to 

0, 0.25, 0.5, or 1.0 ppm O3 showed a concentration-dependent increase in injured cells. Injury 

was evident in proximal and distal conduction airways, lowest in terminal bronchioles, and 

highest in the small side branches downstream of bifurcations. These exposure levels did not 

concurrently elicit changes in LDH activity or total protein in BALF, suggesting that the 

mapping technique is a more sensitive measure of injury and is useful in dosimetry studies. 

P

Whole-body plethysmography of unanesthetized, unrestrained rodents has been used to 

indirectly measure pulmonary resistance (RL) (Shore et al., 2002; Goldsmith et al., 2002; Jang 

et al., 2002). However, these indices of inspiratory/expiratory pressure differences, including 

enh, may be less sensitive than direct measurements of lung airflow resistance (Murphy, 2002). 

Changes in airway structure caused by viral infections also must be considered when evaluating 

laboratory animal studies. Animals with acute viral illness have morphological evidence of 

inflammatory cell infiltration, bronchiolar wall edema, epithelial hyperplasia, and increased 

airway mucous plugs that can cause airway narrowing, air trapping, and serious functional 

changes in the lung (Folkerts et al., 1998). 

Exercise-induced bronchoconstriction in humans appears to be mediated by changes in the 

tonicity of the airway lining fluid (Anderson and Daviskas, 2000). Brannan et al. (1998) suggest 

5-48
that a test in laboratory animals based on the inhalation of mannitol aerosol (hyperosmolar) may 

be feasible and provide information similar to that from exercise challenges in cooperative 

children and adults. Unfortunately, there have been few reports of mannitol or adenosine 

monophosphate challenges in laboratory animals; most studies have utilized histamine, MCh, 

acetylcholine (ACh), or carbachol to determine outcome. In active humans with asthma, 

adenosine monophosphate challenges appear to better reflect ongoing airway inflammation than 

histamine or MCh challenges (Polosa and Holgate, 1997; Avital et al, 1995a,b) and may be 

useful in identifying mechanisms of asthma in laboratory animals and their responsiveness to 

environmental pollutants. 

The increased responsiveness to bronchoconstrictor challenge in asthma is thought to result 

from a combination of structural and physiological factors that include increased inner-wall 

thickness, smooth-muscle responsiveness, and mucus secretion. These factors also are likely to 

determine a level of innate airway responsiveness that is genetically influenced. Chapter 6 

(Section 6.8) discusses cellular and biochemical changes that have been identified in human 

asthmatics. These studies suggest that the mechanisms involved in AHR are multifactoral, with 

general agreement that there is an inconsistent relationship between AHR and markers of 

inflammation. 

A large database of laboratory animal research has been collected on the role of O3 in 

producing an increase in AHR (see Table AX5-12). Exposure levels ($1 ppm for $30 min) 

in many of these studies are not environmentally relevant, but information may be obtained 

regarding the mechanisms of action of O3 concerning: O3 concentration and peak response time, 

inhaled versus intravenous challenge with nonspecific bronchoconstrictors, neurogenic 

mediation, neutrophilic inflammation, and interactions with specific biological agents (e.g., 

antigens and viruses). However, as with other toxicants, high-dose and low-dose mechanisms 

may differ, so interpretation of results must take this into consideration. 

Many species of laboratory animals have been used to study the effects of O3 on airway 

bronchoconstriction. Ozone-induced AHR in guinea pigs has been used to model human 

bronchospasm (van Hoof et al., 1996; 1997a,b; Matsubara et al., 1997a,b; Sun and Chung, 1997; 

Aizawa et al., 1999a,b; Tsai et al., 1998; Nakano et al., 2000). Because these studies were done 

at 2 to 3 ppm O3, these results are not directly relevant for extrapolation to potential airway 

responses in humans exposed to ambient levels of O3. Humans with reactive airway disease 

5-49
(e.g., asthma) appear to be sensitive to ambient levels of O3 (see Chapters 6 and 7) and the 

current understanding is that O3 exacerbates airway responsiveness to specific allergens, 

presumably by nonspecifically increasing AHR. 

Shore et al. (2000, 2002) have shown that O3-induced AHR is reduced in immature rats and 

mice. SD rats exposed to 2 ppm O3 at ages 2, 4, 6, 8, or 12 weeks and A/J mice exposed to 0.3 to 

3 ppm for 3 h at age 2, 4, 8, or 12 weeks had similar concentration-related decreases in V0

E 

0 except at the youngest ages. This smaller decrement in VE suggested a delivered dose that was 

much greater in the younger animals. This group (Shore et al., 2003) has also recently shown 

that obese mice have greater ventilatory responses to O3. Exposures of 2.0 ppm O3 for 3 h to 

lean, WT C57BL/6J and ob/ob mice (mice with a genetic defect in the coding for leptin, the 

satiety hormone) showed that the ob/ob mice had enhanced AHR and inflammation compared to 

the WT mice. These data correlate with epidemiological data showing increased incidence of 

asthma in overweight children. 

Increased AHR to various nonspecific bronchoconstrictive agents (e.g., ACh, MCh, 

histamine, carbachol) given by inhalation or intravenous routes has been previously shown in 

laboratory animals exposed to O3 concentrations #1.0 ppm. Dye et al. (1999) showed 

hyperresponsiveness to MCh in rats 2 h after exposure to 2 ppm O3 for 2 h. AHR can be induced 

by specific antigens as well as O3. The most commonly used laboratory animal model is the 

OVA-sensitized guinea pig. Animals sensitized with OVA have been shown to have similar 

responses to nonspecific bronchoconstrictors as control animals. 

OVA-sensitized guinea pigs (Sun et al., 1997) and mice (Yamauchi et al., 2002) were used 

to determine the enhancement of antigen-induced bronchoconstriction by acute, high-level O3 

(1.0 ppm O3 for 1 h). Male Dunkin-Hartley guinea pigs were sensitized by i.p. injection of OVA 

and exposed to O3 alone, OVA aerosol, or O3 + OVA. Ozone exposure alone increased 

bronchial responsiveness to ACh at 3 h, but not 24 h, whereas OVA alone had no effect. 

Combined exposure to O3 and OVA (1 ppm for 1 h, then 3 min OVA) increased bronchial 

responsiveness to ACh 3 h after O3 exposure. At 24 h following O3 exposure, AHR increased 

when OVA challenge was performed at 21 h, suggesting that O3 preexposure can potentiate 

OVA-induced AHR. Neutrophil counts in the BALF increased at 3 and 24 h after O3 exposure 

alone but were not further increased when O3 exposure was combined with OVA airway 

challenge; however, protein content of the BALF increased at 3 and 24 h in the O3 and OVA 

5-50
groups. Thus, this study also indicates that high-ambient O3 exposure can augment antigen 

(OVA)-induced AHR in guinea pigs. 

Yamauchi et al. (2002) sensitized male C57BL/6 mice by i.p. injection of OVA and then 

exposed them to O3. The sensitized mice had AHR to MCh. Ozone exposure caused significant 

decreases in dynamic lung compliance (C 0 dyn), VE, and PaO2 in OVA-sensitized mice but not in 

controls. A marker of inflammation (soluble intercellular adhesion molecule-1 [sICAM-1]) was 

elevated in the BAL fluid of OVA-sensitized mice, but sICAM-1 levels were not significantly 

changed by O3 exposure, indicating that the O3-induced AHR to MCh was not caused by O3­

induced inflammation. 

Ozone-induced AHR may be temporally associated with inflammatory cells stimulated by 

cytokines (Koto et al., 1997), mast cells (Igarashi et al., 1998; Noviski et al., 1999), or by oxygen 

radicals (Takahashi et al., 1993). One study, however, has shown that inflammation is not a 

prerequisite of AHR (Koto et al., 1997), and it has been suggested that O3-induced AHR may be 

epithelium-dependent (McGraw et al., 2000). For example, neonatal rats pretreated with 

capsaicin, which will permanently destroy C-fibers and prevent O3-induced (1 ppm, 8 h) release 

of neuropeptides (Vesely et al., 1999a), and then exposed to O3 when adults, showed a marked 

increase in airway responsiveness to inhaled aerosolized MCh (Jimba et al., 1995). Takebayashi 

et al. (1998) has shown that depletion of tachykinins by capsaicin treatment, or by a specific 

tachykinin receptor antagonist, can block the induction of AHR by O3. The seemingly disparate 

responses in laboratory animals may be due to species- or strain-specific differences in inherent 

reactivity to bronchoconstrictors or to inherent differences in susceptibility to O3-induced 

inflammation (Zhang et al., 1995; Depuydt et al., 1999; Dye et al., 1999). 

Studies that may be potentially relevant to ambient levels of O3 were conducted in vivo, in 

an isolated perfused lung model, and in ex vivo lung segments using multihour and repeated 

multihour exposures with ambient levels of O3. A study on the relationship between O3-induced 

AHR and tracheal epithelial function was conducted in New Zealand White rabbits by Freed 

et al. (1996). Rabbits exposed to O3 (0.2 ppm for 7 h) demonstrated significantly decreased 

tracheal transepithelial potential difference but no changes in lung resistance. Changes in the 

compartmentalized lung resistance, measured in response to ACh challenge before and after 

bilateral vagotomy, were not significantly different in air-exposed rabbits; however, bilateral 

vagotomy enhanced peripheral lung reactivity in O3-exposed rabbits. The ACh-induced 140% 

5-51
increase in lung resistance with O3 exposure was two times higher than with air exposure, 

indicating that ambient-level O3 exposure affects tracheal epithelial function in rabbits and 

increases central airway reactivity, possibly through vagally-mediated mechanisms. 

Pulmonary mechanics and hemodynamics were studied in the New Zealand White rabbit 

isolated perfused lung model that allowed partitioning of the total pressure gradient into arterial, 

pre- and postcapillary, and venous components (Delaunois et al., 1998). Exposures to O3 

(0.4 ppm for 4 h) were followed by evaluation of airway responsiveness to ACh, substance P 

(SP), or histamine immediately or 48 h later. Ozone inhibited pulmonary mechanical reactivity 

to all three bronchoconstrictors that persisted for 48 h and modified vasoreactivity of the 

vascular bed, but only at 48 h PE. Arterial segmental pressure, normally insensitive to ACh and 

SP, was significantly elevated by O3; precapillary segmental pressure decreased in response to 

Ach, suggesting that O3 can induce direct vascular constriction, but the vascular responses are 

variable and depend on the agonist used and on the species studied. 

Airway responsiveness to the same three compounds was evaluated by Segura et al. (1997) 

in guinea pigs exposed to O3 (0.15, 0.3, 0.6, or 1.2 ppm for 4 h). Ozone did not cause AHR to 

ACh or histamine, except at the highest concentration (1.2 ppm O3) for histamine. However, O3 

caused AHR to SP at $0.3 ppm, suggesting that O3 destroys neutral endopeptidases (responsible 

for SP inactivation) in airway epithelial cells. Vargas et al. (1998), in a follow-up study, 

demonstrated that guinea pigs chronically exposed to 0.3 ppm O3 for 4 h/day became adapted to 

SP-induced AHR. Ozone caused increased sensitivity to SP after 1, 3, 6, 12, and 24 days of 

exposure that was associated with airway inflammation; however, after 48 days of exposure, the 

increased sensitivity to SP was lost. 

This study is in accordance with Szarek et al. (1995) who demonstrated that AHR 

associated with acute O3 exposures does not persist during long-term exposure to near-ambientlevels of O3 (#1 ppm). Fischer 344 rats, exposed to 0.0, 0.12, 0.5, or 1.0 ppm O3, 6 h/day, 

5 days/week for 20 months, demonstrated significantly reduced responses to bethanechol, ACh, 

and electrical field stimulation in eighth generation airway segments. This suggests that some 

adaptation had taken place during long-term exposure, possibly due to increased inner wall 

thickness. 

It is well known that the changes in breathing pattern and lung function caused by O3 are 

attenuated with repeated daily exposures for at least 3 to 5 days. But guinea pigs exposed to 

5-52
0.5 ppm O3, 8 h/day for 7 days showed enhancement of responsiveness of rapidly adapting 

airway receptors (Joad et al., 1998). Repeated exposure increased receptor activity to SP, MCh, 

and hyperinflation; there were no significant effects on baseline or SP- and MCh-induced 

changes in lung compliance and resistance, suggesting that the responsiveness of rapidly 

adapting receptors was enhanced. 

Male and female Hartley guinea pigs exposed to O3 (0.1 and 0.3 ppm, 4 h/day, 4 days/week 

for 24 weeks) were evaluated for airway responsiveness following ACh or OVA inhalation 

challenges (Schlesinger et al., 2002a,b). Ozone exposure did not cause AHR in nonsensitized 

animals but did exacerbate AHR to both ACh and OVA in sensitized animals that persisted for 

4 weeks after exposure. The effects of O3 on airway responsiveness were gender independent 

and were concentration-related for the ACh challenges. 

Schelegle et al. (2003a) evaluated airway responsiveness in infant rhesus monkeys exposed 

to a 5-day O3 episode repeated every 14 days over a 6-month period. Half of the monkeys were 

sensitized to HDMA at 14 and 28 days of age before exposure to a total of 11 episodes of O3 

(0.5 ppm, 8 h/day for 5 days followed by 9 days of FA), HDMA, or O3 + HDMA. Baseline Raw 

was significantly elevated after 10 exposure episodes in the HDMA + O3 group compared to the 

FA, HDMA, and O3 exposure groups. Aerosol challenge with HDMA at the end of the 10th 

episode did not significantly affect Raw, VT, fB, or SaO2. Aerosol challenge with histamine was 

not significantly different after 6 episodes; however, the EC150 Raw for the HDMA + O3 group 

was significantly reduced after 10 episodes when compared to the FA, HDMA, and O3 exposure 

groups, indicating the development of AHR in this group sometime between episodes 6 and 10. 

The results are consistent with altered structural development of the conducting airways. 

During repeated episodic exposures to O3, respiratory responses are first altered to a rapid, 

shallow breathing pattern, which has long been considered protective, especially to the deep 

lung. This dogma has been discounted recently as discussed above (Schelegle et al., 2001). 

Alfaro et al. (2004) examined the site-specific deposition of 18O (1 ppm 2 h) at breathing 

frequencies of 80, 120, 160, or 200 bpm. At all frequencies, parenchymal areas had a lower 

content of 18O than trachea and bronchi. As breathing frequency increased from 80 to 160 bpm, 

the deposition showed a reduction in mid-level trachea and an increase in both mainstream 

bronchi. At this frequency there was also an increase in deposition in parenchyma supplied by 

short (cranial) airway paths, consistent with the results of Schelegle et al., (2001). 

5-53
At 200 bpm, 18O deposition in trachea increased, concurrent with increases in right cranial and 

caudal bronchi regions. Right cranial parenchymal content decreased at 200 bpm, whereas right 

caudal parenchymal levels did not change at any breathing frequency. The authors list some 

limitations of this study, such as the possible effect on regional distribution of ventilation by use 

of the negative-pressure ventilator, the effect of paralysis on airway geometry, and possible 

translocation of 18O during the 2 h exposure period. These two studies provide evidence 

that O3-induced rapid, shallow breathing creates a more evenly distributed injury pattern, 

with possibly greater protection from focal injury to the large conducting airways, including the 

trachea and the left mainstem bronchus. 

Another study of the adaptive phenomena in SD rats used an exposure protocol consisting 

of 5 days of daily 8 h, 1 ppm O3 exposures followed by 9 days of recovery in FA (Schelegle 

et al., 2003b). This O3/FA pattern was repeated for four cycles and demonstrated that the O3­

induced rapid shallow breathing pattern was followed by adaptation that occurred with each 

cycle. However, the release of SP from the trachea, the neutrophil content, and cell proliferation 

became attenuated after the first cycle, suggesting a disconnect from the rapid, shallow breathing 

response. Hypercellularity of the CAR epithelium and thickening of the CAR interstitium, not 

linked to changes in cell proliferation, were also found. The authors suggest mechanism(s) of 

injury from repeated O3 exposures, consisting of diminished neutrophilic inflammation/and or 

release of mitogenic neuropeptides, depressed cell proliferative response, and cumulative distal 

airway lesion. 

Following the initial response of a rapid, shallow breathing pattern, animals eventually 

adapt with continued episodic exposure despite the continued presence of epithelial damage, 

altered structural development, and inflammation of the airways. Chen et al. (2003) used a 

subset of the monkeys from the Schlegele et al. (2003a) study to demonstrate that the attenuation 

of O3-induced rapid, shallow breathing and lung function changes typically seen with 

repeated O3 exposure may be caused by the adaptation of the respiratory motor responses. This 

episodic O3 exposure appeared to create neuroplasticity of the nucleus tractus solitarius (NTS, a 

region of the brainstem which controls respiration), including increased nonspecific excitability 

of the NTS neurons, an increased input resistance, and an increased spiking response to 

intracellular injections of depolarizing current. 

5-54
5.2.5.4 Summary and Conclusions—Effects on Airway Responsiveness 

Ozone-induced AHR has been reported in a number of laboratory species at an exposure 

range of 0.5 to 1.0 ppm and in human asthmatics at ambient levels. In asthmatics, O3 is thought 

to exacerbate AHR to specific allergens by nonspecifically increasing AHR. New studies have 

demonstrated that AHR in asthmatics is due in part to chronic inflammation and airway 

remodeling. Animal studies have shown that O3 exposure can augment OVA-induced AHR. 

Importantly, there is a temporal relationship between inflammatory cell influx and O3-induced 

AHR, but inflammation is not a prerequisite of AHR. Repeated O3 exposures enhance AHR, 

possibly by modulating rapidly adapting airway receptors or by altering the structure of 

conducting airways. 

Currently reported investigations on AHR with repeated O3 exposure to nonsensitized 

laboratory animals have shown equivocal results, especially at the most relevant ambient O3 

concentrations of #0.3 ppm. The few available studies in sensitized laboratory animals are 

consistent with the O3-induced exacerbation of AHR reported in atopic humans with asthma (see 

Chapter 6), but the results are difficult to extrapolate because of interindividual and interspecies 

differences in responsiveness to bronchoprovocation and possible adaptation of airway 

responsiveness with long-term, repeated O3 exposures. Therefore, further studies in laboratory 

animals are needed to investigate responses to the different challenges in relation to 

measurements of airway inflammation and the other physiological and structural factors known 

to contribute to airway responsiveness in human subjects. 

Important new information indicates that rapid, shallow breathing in response to O3 causes 

a more evenly distributed injury pattern rather than protection from injury. New insights into the 

mechanisms of O3-induced AHR suggest that: (1) exercise-induced bronchoconstriction may be 

mediated by changes in tonicity of the bronchial smooth muscles; (2) vagally mediated 

mechanisms may affect tracheal epithelial function and increase central airway reactivity; 

(3) O3 may induce direct vascular constriction; (4) O3 may destroy neural endopeptidases in 

airway epithelial cells, thus preventing the inactivation of SP; and (5) repeated O3 exposures may 

diminish neutrophilic inflammation, depress cell proliferation, and cause cumulative distal 

airway lesions. 

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5.2.6 Genotoxicity Potential of Ozone 

There has been an historical interest in the ability of ground-level pollution to cause cancer, 

especially lung cancer. This interest has been amplified in recent years by results of an 

epidemiologic study that suggest association of increased risks of incident lung cancer with 

elevated long-term ambient concentrations of O3, coarse particulate matter (PM10), and sulfur 

dioxide (SO2) in nonsmoking California males (Beeson et al., 1998; Abbey et al., 1999). 

However, another larger, nationwide American Cancer Society study (Pope et al., 2002) showed 

no significant effect of O3 on mortality risk but showed positive associations between 

warm-season (July-September) O3 concentrations and cardiopulmonary mortality. Studies of 

children and young adults of southwest metropolitan Mexico City, repeatedly exposed to high 

levels of O3, particulate matter (PM), nitrogen oxides (NOx), aldehydes, metals, and other 

components in a complex ambient mixture, also report DNA damage in blood leukocytes and 

nasal epithelial cells (Valverde et al., 1997; Calderón-Garcidueñas et al., 1999) and abnormal 

nasal biopsies (Calderón-Garcidueñas et al., 2001). 

A number of experimental studies have been done to explore the mutagenic/carcinogenic 

potential of O3. In vitro studies are difficult to interpret due to very high exposure levels and 

culture systems that allowed the potential formation of artifacts. Some recently published in 

vivo exposure studies (see Table AX5-13) found increased DNA strand breaks in respiratory 

cells from guinea pigs (Ferng et al., 1997) and mice (Bornholdt et al., 2002) but, again, only with 

exposure to high doses of O3 (1 ppm for 72 h and 1 or 2 ppm for 90 min, respectively). 

Exposing the A/J mouse strain (known to have a high incidence of spontaneous pulmonary 

adenomas) to 0.12, 0.50, and 1.0 ppm O3 for 6 h/day, 5 days/week for up to 9 months, Witschi 

et al. (1999) did not find O3 exposure-related differences in lung tumor multiplicity or incidence. 

Similarly, in a subchronic exposure study (B6C3F1 mice to 0.5 ppm O3 for 6 h/day, 5 days/week 

for 12 weeks), Kim et al. (2001) did not find statistically significant increases in the incidence of 

lung tumors. Significant differences in mean body weight as well as mean absolute and relative 

weights of several organs (e.g., liver, spleen, kidney, testes, and ovary) were observed between 

O3-exposed and air-exposed mice. Histopathologic examination of major organs revealed 

oviductal carcinomas in 3/10 O3-exposed female mice. 

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5.2.6.1 Summary and Conclusions—Genotoxicity Potential of Ozone 

The weight of evidence from new experimental studies does not appear to support 

ambient O3 as a pulmonary carcinogen in laboratory animal models. These new data are in 

agreement with a study of carcinogenicity of O3 from the National Toxicology Program (NTP) 

study (National Toxicology Program, 1994; Boorman et al., 1994), which was negative in male 

and female rats, ambiguous in male mice, and positive only in female mice at high 

concentrations of O3 (i.e., 1.0 ppm). As none of the new experimental studies of genotoxicity 

provided lifetime exposure durations such as those used in NTP cancer studies, the observation 

of no effects must be tempered by consideration of the limited duration of the exposure. Overall, 

then, the new animal studies are inconclusive, as are the epidemiologic studies discussed in 

Chapter 7, which may be due to significant species differences in this health endpoint. Also, O3 

could act as a co-carcinogen functioning to stimulate hyperplasia. In epidemiology studies, 

exposures typically consist of mixtures of copollutants, some of which are known carcinogens 

(see Section 5.4.3). 

5.3 SYSTEMIC EFFECTS OF OZONE EXPOSURE 

Ozone indirectly affects organs beyond the respiratory system due to O3 reaction products 

entering the bloodstream and being transported to target sites. Extrapulmonary effects could 

also be due to the exposure-related production of mediators, metabolic products, and cell 

trafficking. Although systemic effects are of interest and indicate a very broad array 

of O3 effects, they are of limited influence and difficult to interpret. By protecting from 

respiratory tract effects, these systemic effects will likely be protected against also. Systemic 

effects are only summarized briefly here and in Table AX5-14. 

5.3.1 Neurobehavioral Effects 

Animal behavior, both motor activity and operant behavior, has been shown to be 

suppressed by acute O3 exposures (3 to 6 h) of 0.12 ppm. There is a dose dependent decrease in 

activity with increasing exposure levels. Additionally, these lowered activity levels tend to 

attenuate with longer exposure periods. New studies in adult laboratory animals confirm that 

environmentally relevant O3 concentrations from 0.2 to 1.0 ppm can decrease motor activity and 

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affect short- and long-term memory, as tested by passive avoidance conditioning in 4 h 

exposures in rats (Rivas-Arancibia et al., 1998; Avila-Costa et al., 1999; Dorado-Martinez et al., 

2001) or water-maze learning tasks in mice following a 30-day exposure (Sorace et al., 2001). 

The effects have been attributed to reactive oxygen/nitrogen species and/or ozonation products. 

The memory deficits could be blocked by administration of "-tocopherol (Guerrero et al, 1999) 

or taurine (Rivas-Arancibia et al., 2000). Increased freezing and decreased exploratory 

behaviors were accompanied by decreased serotonin levels and increased levels of NO, 

glutamate, dopamine, and striatal lipoperoxidation in rats exposed to 1 ppm of O3 for 4 h (RivasArancibia et al., 2003). The O3-exposed animals also demonstrated neuronal cytoplasm and 

dendrite vacuolation and dilation of rough endoplasmic reticulum cisterns, which the authors 

interpret as a neurodegenerative process resulting from the oxidative stress of acute O3 exposure. 

Niño-Cabrera et al. (2002) demonstrated that a 0.7 ppm O3 exposure for 4 h can induce 

ultrastructural alterations in the hippocampus and prefrontal cortex in aged rats. These are areas 

of the brain where degenerative age-related changes in learning and memory functions have been 

reported (Bimonte et al., 2003). 

Paz (1997) reviewed a series of studies that demonstrated significant alterations of 

electroencephalographic (EEG) patterns during sleep in animals acutely exposed to O3 (0.35 to 

1.0 ppm). Rats and cats both showed loss of paradoxical sleep time after 2 to 8 h of O3 exposure 

(Paz and Bazan-Perkins, 1992; Paz and Huitrón-Reséndiz, 1996). Increased total wakefulness, 

alterations in circadian rhythm, and a permanent 50% loss of paradoxical sleep time were shown 

in rat pups born to dams exposed to 1.0 ppm O3 during gestation (Haro and Paz, 1993). Effects 

on sleep patterns were associated with alterations in brain neurotransmitter levels (HuitrónReséndiz et al., 1994; González-Piña and Paz, 1997) thought to be caused by O3 reaction 

products or prostaglandins (Koyama and Hayaishi, 1994). The permanent effects in pups caused 

by high O3 exposure during gestation were attributed to the diminished antioxidant capability of 

fetal tissue (Günther et al., 1993). 

High, nonambient levels of O3 (e.g., >1.0 ppm) affect visual and olfactory neural pathways 

in the rat. For example, Custodio-Ramierez and Paz (1997) reported a significant delay in visual 

evoked potentials recorded in the visual cortex and the lateral geniculate nucleus of male Wistar 

rats acutely exposed to high levels of O3 (1.5 and 3.0 ppm for 4 h). Colin-Barenque et al. (1999), 

using the same strain, reported cytological and ultrastructural changes in the granular layer of the 

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olfactory bulb after a 4 h exposure to 1 to 1.5 ppm O3. Although these neural effects are thought 

to be caused by O3 reaction products, especially free radicals, the studies do not add much to an 

understanding of the underlying mechanisms. 

5.3.2 Neuroendocrine Effects 

Early studies suggested an interaction of O3 with the pituitary-thyroid-adrenal axis, because 

thyroidectomy, hypophysectomy, and adrenalectomy protected against the lethal effects of high 

concentrations of O3. Concentrations of 0.7 to 1.0 ppm O3 for a 1 day exposure in male rats 

caused changes in the parathyroid; thymic atrophy; decreased serum levels of thyroid stimulating 

hormone, triiodothyronine (T3), thyroxine (T4), free T4, and protein binding; and increased 

prolactin. In more recent studies, increased toxicity to O3 was reported in hyperthyroid rats by 

Huffman et al. (2001) and T3 supplementation was shown to increase metabolic rate and 

pulmonary injury in the lungs of O3-treated animals (Sen et al., 1993). 

The mechanisms by which O3 affects neuroendocrine function are not well understood. 

Cottet-Emard et al. (1997) examined catecholamine activity in rat sympathetic efferents and 

brain areas of prime importance to adaptation to environmental stressors. Exposures of 

0.5 ppm O3 for 5 days caused inhibition of norepinephrine turnover in heart (!48% of the 

control level) but not in lungs and failed to modify the tyrosine hydroxylase activity in superior 

cervical ganglia and the catecholamine content in the adrenal glands. In the CNS, O3 inhibited 

tyrosine hydroxylase activity in noradrenergic brainstem cell groups and decreased 

catecholamine turnover in the cortex (!49%) and striatum (!18%) but not in the hypothalamus. 

This suggests that high ambient levels of O3 can produce marked neural disturbances in 

structures involved in the integration of chemosensory inputs, arousal, and motor control, effects 

that may be responsible for some of the behavioral effects seen with O3 exposure. 

5.3.3 Cardiovascular Effects 

Studies of the effects on hematological parameters and blood chemistry in rats have shown 

that erythrocytes are a target of O3. Exposures to 1.0 ppm O3 for 3 h have been found to 

decrease heart rate (HR), mean arterial pressure (MAP), and core temperature (TCO) and to 

induce arrhythmias with some exposures in rats. These effects are more pronounced in adult and 

awake rats than in younger or sleeping animals. Exposures of 0.2 ppm for 48 h have been shown 

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to cause bradycardia, while exposures of 0.1 ppm O3 for 3 days have been shown to cause 

bradyarrhythmia in these animals. 

A more recent study of rats exposed to FA for 6 h, followed 2 days later by a 5 h exposure 

to 0.1 ppm O3, 5 days later by a 5 h exposure to 0.3 ppm O3, and 10 days later by a 5 h exposure 

to 0.5 ppm O3, used the head-out plethysmograph for continuous measurements (Arito et al., 

1997). Each of the O3 exposures was preceded by a 1 h exposure to FA. Transient rapid, 

shallow breathing with slightly increased HR appeared 1 to 2 min after the start of O3 exposures 

and was attributed to an olfactory response. Persistent rapid, shallow breathing with a 

progressive decrease in HR occurred with a latent period of 12 h. During the last 90-min of 

0 exposure, averaged values for relative VE tended to decrease with the increase in O3 

concentration for young (4 to 6 months) but not old (20 to 22 months) rats. 

Studies by Watkinson et al. (1995, 2001) and Highfill and Watkinson (1996), that utilized 

radiotelemetry transmitters in unanesthetized unrestrained rats, demonstrated that when HR was 

reduced during a 5-day, 0.5 ppm O3 exposure, TCO and activity levels also decreased. The 

decreases in TCO and BP reported in these studies and by Arito et al. (1997) suggest that the 

changes in ventilation and HR are mediated through physiological and behavioral defense 

mechanisms in an attempt to minimize the irritant effects of O3 inhalation. Decreased activity 

was previously reported in laboratory animals during exposure to O3 (see above). 

Similar cardiovascular and thermoregulatory responses in rats to O3 were reported by 

Iwasaki et al. (1998). Repeated exposure to 0.1, 0.3, and 0.5 ppm O3 8 h/day for 4 consecutive 

days caused disruption of circadian rhythms of HR and TCO on the first and second exposure 

days that was concentration-dependent. The decreased HR and TCO recovered to control values 

on the third and fourth days of O3 exposure. 

The thermoregulatory response to O3 was further characterized by Watkinson et al. (2003). 

Male Fischer-344 rats were either exposed to 0.0 ppm for 24 h/day (air), 0.5 ppm for 6 h/day 

(intermittent), or to 0.5 ppm for 23 h/day (continuous) at 3 temperatures, 10 °C (cold), 22 °C 

(room), or 34 °C (warm). Another protocol examined the effects of O3 exposure (0.5 ppm) and 

exercise (described as rest, moderate, or heavy) or CO2-stimulated ventilation. Both intermittent 

and continuous O3 exposure caused decreases in HR and TCO and increases in BALF 

inflammatory markers. Exercise in FA caused increases in HR and TCO while exercise in O3 

caused decreases in those parameters. Carbon dioxide and O3 induced the greatest deficits in HR 

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and TCO. Several factors were suggested that may modulate the hypothermic response, including 

dose, animal mass, and environmental stress. 

Laboratory animals exposed to relatively high O3 concentrations ($0.5 ppm) demonstrate 

tissue edema in the heart and lungs. This may be due to increased circulating levels of atrial 

natriuretic factor (ANF), which is known to mediate capillary permeability, vasodilation, and BP 

(Daly et al., 2002). Increased levels of ANF were reported in the heart, lungs, and circulation of 

rats exposed to 0.5 ppm O3 for 8 h ( Vesely et al., 1994a,b,c). 

Earlier work demonstrated O3-induced release of functionally active PAF from rodent 

epithelial cells and the presence of PAF receptors on AMs. New work examining lipid 

metabolism (Section 5.2.1.4) and mediators of inflammatory response and injury 

(Section 5.2.3.4) confirm earlier findings indicating that PAF (Kafoury et al., 1999) and PAF 

receptors (Longphre et al., 1999) are involved in responses to O3. In addition to the role of PAF 

in pulmonary inflammation and hyperpermeability, this potent inflammatory mediator may have 

clotting and thrombolytic effects, though this has not been demonstrated experimentally (see 

Figure 5-2). This cardiovascular effect may help explain, in part, some limited epidemiologic 

findings suggestive of possible association of heart attack and stroke with ambient O3 exposure 

(see Chapter 7). The findings of Pulfer and Murphy (2004) and Pulfer et al. (2005), as discussed 

in Section 5.2.1.4, which characterize the in vitro and in vivo production of two biologically 

active oxysterols, are also suggestive of a mechanism whereby O3 exposure might be implicated 

in the increased risk of cardiopulmonary disease. 

5.3.4 Reproductive and Developmental Effects 

Early studies of pre- and postnatal exposure to O3 were performed at relatively high 

concentrations. Teratogenic effects were not observed with intermittent exposures of 0.44 to 

1.97 ppm O3 during any part of gestation. Continuous exposure during mid-gestation increased 

the resorption of embryos while exposures during late gestation delayed some behavioral 

developments (e.g., righting, eye opening). There were no effects on neonatal mortality up to 

1.5 ppm O3, whereas some transient effects on weight gain were observed at exposures of 

0.6 ppm O3. 

More recent studies tend to confirm previous conclusions that prenatal exposures to O3 

concentrations <1.0 ppm do not cause major or widespread somatic or neurobehavioral effects in 

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the offspring of laboratory animals. These studies generally add some weight toward a negative 

interpretation of the importance of contributions of low, ambient O3 to lower birth weights and 

gross development defects reported in neonates born to women exposed to typical ambient 

pollution (e.g., Renner, 2002; Chen et al., 2002; Ritz and Yu, 1999). Some postnatal O3 

exposure studies continue to find a few subtle or borderline somatic and behavioral deficits that 

will require further research to better assess potential risk to developing humans. 

Studies of somatic and neurobehavioral development in female CD-1 mice exposed during 

pregnancy (days 7 to 17) to O3 (0, 0.4, 0.8, or 1.2 ppm) failed to show any O3 effects on 

reproductive or behavioral performance (Bignami et al., 1994). The study did find significant 

decreases in body weight gain and delayed eye opening in pups in the 1.2 ppm exposure group. 

The lack of effect on behavioral performance contrasts with earlier findings, which may be due 

to the use of different species, differing exposure durations, cross-fostering used in the latter 

study, different species, and exposure durations during pregnancy. A second study using CD-1 

mice exposed in utero from conception through day 17 of pregnancy to 0, 0.2, 0.4, and 

0.6 ppm O3 found no significant deficits in reproductive performance, postnatal somatic and 

neurobehavioral development, or adult motor activity (Petruzzi et al., 1995). A third study by 

the same group (Petruzzi et al., 1999), using O3 exposures (0.3, 0.6, or 0.9 ppm) that continued 

postnatally until weaning, showed subtle changes in handedness and morphine reactivity. 

Exposures to 0.6 ppm O3 caused a reduced preference for the right paw in adulthood. Exposures 

to 0.9 ppm O3 altered hot plate avoidance after i.p. treatment with morphine in adulthood. 

CD-1 mice exposed to 0.6 ppm O3 from birth through weaning demonstrated no 

impairment of navigational performance during acquisition and only subtle changes during 

reversal (Dell’Omo et al., 1995a). Additionally, there were no O3-induced effects on 

reproductive performance, but offspring showed a significant reduction in body weight. Effects 

on neurobehavioral development with this exposure were minor, with some attenuation of 

activity responses and impairment of passive avoidance acquisition (Dell’Omo et al. (1995b). 

The offspring of CD-1 mice continuously exposed from 30 days prior to the formation of 

breeding pairs until postnatal day 17 to 0.0, 0.3, or 0.6 ppm O3 showed only small and selective 

effects on somatic and sensorimotor development (Sorace et al., 2001). 

Morphological changes were found in the anterior cerebellar lobe of rat pups born to dams 

exposed during the entire gestation period to very high (1.0 ppm) O3 concentrations for 12 h/day 

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(Rivas-Manzano and Paz, 1999). Additionally, the dams displayed significantly fewer 

implantations, increased rate of reabsorptions, a high incidence of spontaneous abortion, and 

offspring with low birth weight, as noted by previous investigators. 

5.3.5 Effects on the Liver, Spleen, and Thymus 

Early investigations of the effects of O3 on liver centered on xenobiotic metabolism, and 

the prolongation of sleeping time, which was observed at 0.1 ppm O3. In some species, only 

adults and especially females were affected. In rats, high (1.0 to 2.0 ppm for 3 h) acute O3 

exposures caused increased production of NO by hepatocytes and enhanced protein synthesis 

(Laskin et al., 1994; 1996). The O3-associated effects shown in the liver are thought to be 

mediated by inflammatory cytokines or other cytotoxic mediators released by activated 

macrophages in the lungs (Vincent et al., 1996; Laskin et al., 1998; Laskin and Laskin, 2001). 

Except for the earlier work on xenobiotic metabolism, the responses occurred only after very 

high acute O3 exposures. 

Examinations of the effects of O3 on spleen and thymus have shown that O3 primarily 

affects T-cell mediated systemic immunity. As with the O3-associated effects shown in the liver, 

most of the statistically significant changes occurred after acute exposures to very high O3 

concentrations and relate to systemic oxidative stress. Using more relevant ambient urban O3 

exposure patterns, effects were not found on systemic immune function of rats. 

5.3.6 Effects on Cutaneous and Ocular Tissues 

Ozone exposure not only affects various organ systems, when inhaled, but also has direct 

effects on the exposed skin and eyes. The outermost layer of the skin (SC, stratum corneum) 

may be oxidized, which can lead to compromise of the skin barrier and an epidermal proinflammatory response ( Weber et al., 2001; Thiele, 2001). These effects are found only at very 

high concentrations (>1 to 5 ppm) and have not been shown at more relevant ambient levels of 

exposure. The skin possesses a well-developed defense system against oxidative stress, utilizing 

nonenzymatic (e.g., vitamin C, GSH, UA, "-tocopherol) and enzymatic (e.g., SOD, catalase, GR, 

and GSHPx) antioxidants (Cross et al., 1998). Ocular tissues have similar antioxidant protective 

functions as the skin but are not as well studied (Mucke, 1996; Rose et al., 1998). Effects of 

ground-level smog on the eyes have been reported but generally are attributed to related 

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photochemical oxidants like peroxyacetyl nitrate (PAN) (Vyskocil et al., 1998) or possibly to 

atmospheric O3 precursors or reaction products like aldehydes. As in other tissues, O3 may have 

disparate high-dose and low-dose mechanisms of effect on skin and eyes; so, results must be 

interpreted in this light. 

Hairless mice (SKH-1) exposed to O3 (0.8 to 10 ppm for 2 h) were used to demonstrate 

that O3 depletes the low molecular weight antioxidants (e.g., "-tocopherol, vitamin C, GSH, UA) 

in the SC at $1.0 ppm and causes increased MDA at $5 ppm (Weber et al., 1999, 2000, 2001). 

Valacchi et al. (2000) demonstrated that preexposure to 0.5 O3 for 2 h followed by low-dose 

ultraviolet (UV) radiation (0.33 MED) caused depletion of "-tocopherol. This suggests that 

combined low doses of UV radiation and near-ambient levels of O3 may cause oxidative stress 

on the SC. Prolonged exposure to 0.8 ppm O2 for 6 h also induces cellular stress responses that 

included the formation of HNE protein adducts, HSP27, and heme-oxygenase-1 in the deeper 

cellular layers of the skin that continued for up to 18 h after O3 exposure, followed by repair 

processes (Valacchi et al., 2003). 

The importance of O3 and UV-induced cellular protein oxidation found in murine skin 

models to possibly similar environmentally-induced changes in human SC keratins was 

identified by Thiele et al. (1998, 1999) and Thiele (2001). Using the presence of carbonyl 

groups in proteins as a marker of reactive oxygen-mediated protein oxidation, they reported 

higher carbonyl levels in the upper SC from the tanned skin of humans and in the skin of healthy 

human volunteers exposed to model chemical oxidants (e.g., hypochlorite, benzoyl peroxide) 

that were inversely correlated with "-tocopherol levels. The environmentally-induced oxidative 

damage identified in human SC represents an early pathophysiological stage in the development 

of barrier disruption and inflammation, and possibly has implications for the process of 

desquamation. The relevance of potentiation of environmental oxidative stress by O3 exposure 

of human skin needs further study. 

5.3.7 Summary and Conclusions—Systemic Effects of Ozone 

Neurobehavioral effects of O3 at concentrations of 0.2 to 1.0 ppm include decreased motor 

activity, short- and long-term memory deficits, increased freezing behavior, and decreased 

exploratory behaviors. These effects have been associated with reactive oxygen/nitrogen 

species, ozonation products, altered neurotransmitter levels, morphological changes in several 

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brain regions, and altered EEG patterns during sleep. Neuroendocrine effects of O3 include 

morphological and hormonal changes in the pituitary-thyroid-adrenal axis at concentrations 

of ~0.75 ppm and alterations of visual and olfactory neural pathways at concentrations >1 ppm. 

Mechanisms underlying these effects are not understood at this time. Cardiovascular effects 

of O3 at concentrations of 0.3 to 0.5 ppm include decreased HR, TCO, and BP, which have been 

termed a hypothermic response. Concentrations of O3 $0.5 ppm cause tissue edema (possibly 

mediated by ANF). 

Prenatal exposures to O3 concentrations <1.0 ppm did not cause noticeable somatic or 

neurobehavioral effects in offspring, while concentrations of 1.0 to 1.5 ppm caused varying 

effects on neonatal mortality. Some studies have shown an effect of O3 on liver xenobiotic 

enzymes at concentrations as low as 0.1 ppm, while other studies have shown no alterations in 

metabolic enzymes at even 1 ppm, with the effects appearing to be highly-species specific. 

Effects on spleen and thymus appear to only occur at high O3 concentrations (>1.0 ppm), while 

relevant ambient urban exposures have no effect on systemic immune function in rats. Effects 

of O3 on cutaneous and ocular tissue are only seen at high, nonrelevant concentrations. 

5.4 INTERACTIONS OF OZONE WITH OTHER CO-OCCURRING 

POLLUTANTS 

Ozone is part of a complex mixture of air pollutants with a composition and pattern that 

varies geographically and temporally (by hour of the day, day of the week, and season). Health 

effects caused by the complex mixture are undoubtedly different (either subtly or significantly) 

from the additive effects of a few of the hundreds of compounds present. The only disciplinary 

approach that can evaluate a “real-world” complex mixture is epidemiology (Chapter 7). Still, 

because of the difficulty in evaluating causative factors and quantitative relationships in 

epidemiology studies, it is useful to consider animal toxicological studies of mixtures. Such 

studies can be divided into three categories: (1) ambient air mixtures, (2) laboratory-generated 

complex mixtures (e.g., gasoline combustion mixtures having UV-irradiation, other reaction 

mixtures with O3 and several other components), and (3) binary mixtures. In most cases, 

experimental designs in the first two classes did not have an O3-only group, making it difficult or 

impossible to discern specifically the influence of O3 per se. The more recent mixture studies 

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that are discussed here typically have been with NO2, sulfuric acid (H2SO4), or ammonium 

sulfate ((NH4)2SO4). 

Interpreting the mixture studies in terms of real-world risk is difficult because laboratory 

exposure patterns do not always represent real-world exposure patterns. As shall be seen, all 

interaction possibilities have occurred, depending upon the composition of the mixture, the 

endpoint examined, and the exposure regimen. In some cases, no interaction was found. Most 

often, additivity (the effects of the mixture are equal to the sum of the effects of the individual 

components) or synergism (the effects of the mixture are greater than the sum of the effects of 

the individual components) was observed. Antagonism (the effects of the mixture are less than 

the sum of the individual components) was rarely found. 

5.4.1 Ozone and Nitrogen Oxides 

The most commonly studied copollutant in binary mixtures with O3 is NO2. Both earlier 

studies and more recent work indicate that, although interaction may occur between these two 

pollutants, in general, O3 often masked the effects of the NO2 or accounted for most of the 

response, due to the greater toxicity of O3. Very generally, additivity occurred after acute 

exposure and synergism occurred with prolonged exposure. Interpreting the O3/NO2 mixture 

studies is challenging, because laboratory exposure patterns rarely simulate real-world exposure 

patterns. NO2 typically peaks before O3, with a mixture occurring between the individual gas 

peaks, but most laboratory exposures used mixtures only. Also, most studies of O3 and NO2 

mixtures used ambient levels of O3 and levels of NO2 high above ambient. Table AX5-15 lists 

more recent studies evaluating coexposures to NO2 and O3. 

Chronic exposures of rats to O3 (0.8 ppm) and NO2 (14.4 ppm) for 6 h/day caused 

development of respiratory insufficiency and severe weight loss. Half of these animals died after 

55 to 78 days of exposure due to severe fibrosis (Farman et al., 1997). Increased total lung 

collagen and elastin were observed, with loss of mature collagen, suggesting breakdown and 

remodeling of the lung parenchyma. Morphological examination following these coexposures 

demonstrates a sequence of events starting with increasing inflammatory and mild fibrotic 

changes for the first 3 weeks of exposure,

 stabilized or even reduced changes after 4 to 6 weeks, 

and severe increases over 7 to 9 weeks of exposure (Farman et al., 1999). This suggests that 

repair processes occurring during the middle 4 to 6 weeks of exposure become overwhelmed, 

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leading to progressive fibrosis after 7 to 8 weeks of exposure. When the coexposure was 

extended for 90 days, lesions were noted far into the acinus, but the extent of tissue involvement 

was the same after 7, 78, and 90 days of exposure. At the end of exposure, high levels of 

procollagen types I and III mRNA were observed within central acini in the lungs from the 

combined exposure group but not in lungs from the rats exposed to O3 or NO2 alone. 

Sprague-Dawley rats exposed to 0.3 ppm O3 and the combined exposure of O3 and 

1.2 ppm NO2 for 3 days demonstrated significant DNA single-strand breaks in AMs (Bermúdez 

et al., 1999). No changes were caused by NO2-only exposure. The same exposures stimulated 

the activity of poly (adenosinediphosphate-ribose) (polyADPR) synthetase, suggesting a 

response to lung cellular DNA repair caused by oxidant-induced lung injury (Bermúdez, 2001). 

The laboratory animal model of progressive pulmonary fibrosis, utilizing long-term combined O3 

(0.4 to 0.8 ppm) and high-level NO2 (7 to 14 ppm) exposure, causes an initial acute pulmonary 

inflammation, followed by adaptation and repair, and eventually causing pulmonary fibrosis 

after 6 to 13 weeks of exposure (Ishii et al., 2000a; Weller et al., 2000). Unfortunately, this 

model is not very useful for understanding potential interactive effects of ambient concentrations 

of O3 and NO2. 

5.4.2 Ozone and Other Copollutants 

Ozone and Formaldehyde 

Early studies with combined exposures to O3 and formaldehyde (HCHO) found evidence 

of both synergistic and non-interactive effects. Newer work listed in Table AX5-16 includes 

studies of biochemical and histopathological endpoints in rats exposed to 0.4 ppm O3 and 

3.6 ppm HCHO, alone and combined, for 8 h/day for 3 days (Cassee and Feron, 1994). They 

demonstrated no interactive effects in the nasal respiratory epithelium, despite the high levels of 

HCHO when compared to typical ambient levels of 1 to 10 ppb (e.g., Rehle et al., 2001). Mautz 

(2003) studied changes in breathing pattern and epithelial cell proliferation using exposures of 

0.6 ppm O3 and 10 ppm HCHO alone and in combination for 3 h with exercise at two times 

resting ventilation. Even with exercise, HCHO does not substantially penetrate to the lower 

respiratory tract to interact with O3 and does not alter breathing patterns to modify local O3 dose. 

Parenchymal injury was, therefore, due to O3 alone. In the nasal transitional epithelium and in 

the trachea, however, combined exposure produced additive effects due to the increased volume 

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of toxicants during exercise. No other combined pollutant studies have been published in the 

peer-reviewed literature, although two studies compared the respiratory effects of O3 to HCHO. 

Nielsen et al. (1999) compared upper airway sensory irritation caused by HCHO concentrations 

up to 4 ppm to the lower airway irritation caused by O3. Using BALB/c mice, they continuously 

measured fB, VT, expiratory flow, time of inspiration, time of expiration, and respiratory patterns 

during acute, 30-min exposures. They reported a no-effect level of 0.3 ppm for HCHO and 

1.0 ppm for O3. 

Thus, O3 and HCHO do not appear to have additive effects except during exercise and that 

is due to an increased volume of gas reaching the tissue. Any possible synergism occurs in the 

nasal epithelium. HCHO exerts its effects primarily in the upper respiratory tract, whereas the 

primary site of acute cell injury from O3 occurs in the conducting airways. EPA is currently 

completing a toxicological and epidemiological review and risk characterization for HCHO. 

Ozone and Tobacco Smoke 

Early studies of combined exposures of O3 (1 ppm) and tobacco smoke demonstrated 

altered airway responsiveness to inhaled bronchoconstrictor challenge and tracheal vascular 

permeability in guinea pigs. Table AX5-17 lists studies completed since the 1996 AQCD which 

evaluated tobacco smoke/O3 coexposures. 

Wu et al. (1997) reported that inhalation of cigarette smoke evokes a transient 

bronchoconstrictive effect in anesthetized guinea pigs. Total pulmonary resistance and Cdyn were 

compared before and after acute exposure to 1.5 ppm O3 for 1 h. Cigarette smoke alone (7 mL) 

at a low concentration (33%) induced a mild and reproducible bronchoconstriction that slowly 

developed and reached its peak after a delay of >1 min. After O3 exposure, the same cigarette 

smoke inhalation challenge evoked an intense bronchoconstriction that occurred more rapidly, 

reaching its peak within 20 s, and being sustained for >2 min. Pretreatment with selective 

antagonists of neurokinin type 1 and 2 receptors completely blocked the enhanced airway 

responsiveness suggesting that O3 exposure induced AHR to inhaled cigarette smoke, primarily 

from the bronchoconstrictive effect of endogenous tachykinins. 

The above studies were conducted with undiluted tobacco smoke and high O3 

concentrations. To determine the effects of aged and diluted sidestream cigarette smoke (ADSS) 

as a surrogate of environmental tobacco smoke (ETS) on O3-induced lung injury, Yu et al. 

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(2002) exposed male B6C3F1 mice to (1) FA, (2) ADSS, (3) O3, or (4) ADSS followed by O3 

(ADSS/O3). Exposure to 30 mg/m3 ADSS, 6 h/day for 3 days, followed by exposure to 

0.5 ppm O3 for 24 h was associated with a significant increase in the number of cells recovered 

by BAL compared with exposure to ADSS alone or O3 alone. Neutrophils, lymphocytes, and 

total protein levels in BAL were increased following the combined exposure when compared 

with all other groups. Within the CAR, the percentage of proliferating cells was unchanged from 

control following exposure to ADSS alone but was significantly elevated following exposure 

to O3 and further augmented in a statistically significant manner in mice exposed to ADSS/O3. 

Following exposure to O3 alone or ADSS/O3, the ability of AMs to release IL-6 under LPS 

stimulation was significantly decreased, while exposure to ADSS alone or ADSS/O3 caused a 

significantly increased release of TNF-" from AMs under LPS stimulation. These data suggest 

that ADSS exposure enhances the sensitivity of animals to O3-induced lung injury. 

Acute exposure to ETS also may make a healthy person more susceptible to sequential O3 

exposure by affecting lung barrier function or the underlying epithelium. Toxicological studies 

with components of ETS (e.g., nicotine receptor agonists, acrolein, and oxidants) have shown 

that the vagal bronchopulmonary C-fibers are stimulated by acute exposures that initiate both 

central and local responses (Bonham et al., 2001; Mutoh et al., 2000). The central responses 

(e.g., tachypnea, cough, bronchoconstriction, increased mucous secretion) are more protective of 

the lungs; however, local responses may include increased sensitization of the C-fibers to other 

irritants, including O3. Active tobacco smokers should not be similarly affected, because they 

already have significant chronic airway inflammation and increased mucus production. In fact, 

chronic smokers appear to have diminished lung function responses to O3 (see Chapter 6). 

5.4.3 Complex (Multicomponent) Mixtures Containing Ozone 

Ambient pollution in most areas is a complex mix of more than two chemicals. A number 

of new studies have examined the effects of exposure to multicomponent atmospheres 

containing O3. Some of these studies attempted to simulate photochemical reaction products 

occurring under actual atmospheric conditions. However, the results of these studies are often 

difficult to interpret because of chemical interactions between the components, as well as the 

resultant production of variable amounts of numerous secondary reaction products, and a lack of 

precise control over the ultimate composition of the exposure environment. In addition, the role 

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of O3 in the observed biological responses is often obscure. Prior studies using irradiated 

automobile exhaust mixtures containing total oxidant concentrations (expressed as O3) in the 

range of 0.2 to 1.0 ppm have demonstrated pulmonary function changes in several species. 

A more recent attempt has been made to examine multicomponent mixtures resulting from 

the reaction of O3 with unsaturated hydrocarbons [e.g., isoprene (C5H8) and terpene (C10H16)], 

which produces HCHO, formic acid, acetone, acrolein, acetic acid, and other oxidation products 

(many of which are strong airway irritants). Wilkins et al. (2001) evaluated sensory irritation by 

measuring mean fB in the mouse bioassay and found a 50% reduction after 30 min of exposure to 

reaction products of O3 and isoprene. The mixture at this time period contained <0.2 ppm O3, so 

the authors attributed the observed effects to the oxidation products. Clausen et al. (2001), using 

the same mouse model, evaluated the reaction products of O3 and limonene. A 33% reduction in 

mean fB was produced after 30 min of exposure to the mixture containing <0.3 ppm O3, again 

implicating the effects of strong irritant products. Further work needs to be done with these 

complex reaction mixtures, because of their potential impact on the respiratory tract. The results 

would be particularly important, however, to the reaction of O3 indoors (see Chapter 3). 

Pollutant mixtures containing acid aerosols comprise another type of commonly examined 

exposure atmosphere (studies summarized in Table AX5-18). Earlier studies that employed 

simultaneous single, repeated, or continuous exposures of various animal species to mixtures of 

acid sulfates and O3 found responses for several endpoints, including tracheobronchial 

mucociliary clearance, alveolar clearance, pulmonary mechanics, and lung morphology, to be 

due solely to O3. Some synergism was noted for bacterial infectivity, response to antigen, and 

effects on lung protein content and the rate of collagen synthesis. 

More recent studies found some differences in airway responses to inhaled acid particle-O3 

mixtures that may have been partly due to airway dosimetry. Various physical and chemical 

mechanisms may be responsible (see Schlesinger, 1995). For example, physical adsorption or 

absorption of O3 or its reaction products on a particle could result in transport to more sensitive 

sites, or to sites where O3, by itself, would not normally be reactive (Madden et al., 2000). 

Chemical reactions on the surface of particles can form secondary products that are more 

toxicologically active, or chemical characteristics of the particle may change the residence time 

or reactivity of oxidation products at the site of deposition. The hypothesis that synergism 

between O3 and sulfates is due to decreased pH changing the residence time or reactivity of 

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reactants, such as free radicals, was tested by Chen et al. (1995) and El-Fawal et al. (1995). 

Male New Zealand White rabbits were exposed for 3 h to 125 μg/m3 H2SO4, 0.1, 0.3, or 

0.6 ppm O3, and to combinations. Chen et al. (1995) demonstrated that decreased pH following 

exposure to acid aerosol was correlated with phagocytic activity and capacity of harvested 

macrophages and that exposure to O3/H2SO4 removed this relationship. El-Fawal et al. (1995) 

showed that responsiveness of rabbit harvested bronchial rings to ACh was increased following a 

3 h O3 exposure, but that 0.1 to 0.6 ppm O3/0.5 to 0.125 mg/m3

 H2SO4 combinations resulted in 

antagonism. 

As discussed in Section 5.2.2.1, Churg et al. (1996) demonstrated increased uptake of 

asbestos or TiO2 in response to 10 min O3 (up to 1.0 ppm) preexposure, suggesting that low 

concentrations of O3 may increase the penetration of some types of PM into epithelial cells. 

Using human epithelial cell cultures, Madden et al. (2000) demonstrated a greater potency for 

ozonized diesel PM to induce PGE2 production. This suggests that 0.1 ppm O3 for 24 h can 

modify the biological activity of PM derived from diesel exhaust. 

Effects of combined exposures of O3 and resuspended urban particles on cell proliferation 

in epithelial cells of the terminal bronchioles and the alveolar ducts were examined by Vincent 

et al. (1997) and Adamson et al. (1999). Rats exposed to 0.8 ppm O3 in combination with 5 or 

50 mg/m3

 particles for 4 h demonstrated greatly potentiated proliferative effects compared to O3 

exposure alone. These findings using resuspended dusts, although at high concentrations, are 

consistent with the studies demonstrating interaction between H2SO4 aerosols and O3. Effects of 

acute coexposure to 0.6 ppm O3 and fine or ultrafine H2SO4 (0.5 to 0.3 mg/m3

) aerosols on lung 

morphology were examined by Kimmel et al. (1997). They demonstrated that alveolar septal 

volume was increased in animals coexposed to O3 and ultrafine, but not fine, H2SO4. 

Interestingly, cell proliferation was increased only in animals coexposed to fine H2SO4 and O3, 

as compared to animals exposed to O3 alone. Subchronic exposure to acid aerosols (20 to 

150 μg/m3

 H2SO4) had no interactive effect on the biochemical and morphometric changes 

produced by either intermittent or continuous exposure to 0.12 to 0.2 ppm O3 for up to 90 days, 

which suggests that the interactive effects of O3 and acid aerosol coexposure in the lung 

disappeared during the long-term exposure (Last and Pinkerton, 1997). Sindhu et al. (1998) 

observed an increase in rat lung putrescine levels after repeated, combined exposures to O3 and a 

nitric acid (HNO3) vapor for 40 weeks. 

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Other studies have examined interactions between carbon particles and O3. The 

interactions of intratracheally instilled carbon particles followed by either a 7-day or 60-day 

exposure to 0.5 ppm O3 in rats was evaluated by Creutzenberg et al. (1995). The carbon 

particles caused diminished phagocytotic capacity and chemotactic migration capability of AMs 

that was stimulated by the subsequent O3 exposure. Inflammatory responses following 

exposures to low- and high-concentration mixtures of O3 and acidic aerosols (0.2 ppm O3 + 

50 μg/m3

 carbon + 100 μg/m3 H2SO4; 0.4 ppm O3 + 250 μg/m3 carbon + 500 μg/m3 H2SO4, 

respectively) for 1 or 5 days was examined by Kleinman et al. (1999). The response with 

the O3-particle mixture was greater after 5 days (4 h/day) than after day 1. This contrasted 

with O3 exposure alone (0.4 ppm), which caused marked inflammation on acute exposure, 

but no inflammation after 5 consecutive days of exposure. 

The effects of a mixture of elemental carbon particles, 0.2 ppm O3, and 0.5 mg/m3 

ammonium bisulfate on rat lung collagen content and macrophage activity was examined by 

Kleinman et al. (2000). Decreases in lung collagen, and increases in macrophage respiratory 

burst and phagocytosis were observed relative to other pollutant combinations. Mautz et al. 

(2001) used a similar mixture (i.e., elemental carbon particles, 0.16 to 0.59 ppm O3, ammonium 

bisulfate 0.5 to 0.22 mg/m3

, but with 0.11 to 0.39 ppm NO2 also) and exposure regimen as 

Kleinman et al. (2000). Also observed were decreases in pulmonary macrophage Fc-receptor 

binding and phagocytosis and increases in acid phosphatase staining. Bronchoalveolar epithelial 

permeability and cell proliferation were increased. Altered breathing-patterns also were 

observed, with some adaptations occurring. 

Bolarin et al. (1997) exposed rats to 50 or 100 μg/m3

 carbon particles in combination with 

ammonium bisulfate and 0.2 ppm O3. Despite 4 weeks of exposure, they observed no changes in 

protein concentration in BALF or in blood prolyl 4-hydroxylase, an enzyme involved in collagen 

metabolism. Slight decreases in plasma fibronectin were present in animals exposed to the 

combined pollutants versus O3 alone. Thus, the potential for adverse effects in the lungs of 

animals challenged with a combined exposure to particles and gaseous pollutants is dependent 

on numerous factors, including the gaseous copollutant, concentration, and time. 

In a complex series of studies, Oberdörster and colleagues examined the interaction of 

several pulmonary oxidative stress pollutants. Elder et al. (2000a,b) reported the results of 

combined exposure to ultrafine carbon particles (100 μg/m3

) and O3 (1 ppm for 6 h) in young 

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and old Fischer 344 rats that were pretreated with aerosolized endotoxin. In old rats, exposure to 

carbon and O3 produced an interaction that resulted in a greater influx in neutrophils than that 

produced by either agent alone. This interaction was not seen in young rats. Oxidant release 

from lavage fluid cells also was assessed and the combination of endotoxin, carbon particles, 

and O3 produced an increase in oxidant release in old rats. This mixture produced the opposite 

response in the cells recovered from the lungs of the young rats, indicating that the lungs of the 

aged animals underwent greater oxidative stress in response to a complex pollutant mix of 

particles, O3, and a biogenic agent. Johnston et al. (2000a, 2002) reported the results of 

combined exposure to O3 (1.0 and 2.5 ppm for 4, 20, or 24 h) and low-dose endotoxin, or to O3 

and endotoxin separately, in newborn and adult C57BL/6J mice. In the first study, adult 

(8 weeks old) mice showed greater sensitivity to O3 than newborn (36 h old) mice on the basis of 

mRNAs encoding for various chemokines and cytokines. In contrast, adult and newborn mice 

responded similarly 2 h after endotoxin exposure (10 ng for 10 min), suggesting that age 

differences in O3-generated inflammation is secondary to epithelial cell injury. In the second 

study, 8-week-old mice exposed to O3 (1 ppm for 24 h) followed by endotoxin (37.5 EU for 

10 min) showed increased responsiveness over either exposure alone, on the basis of increased 

expression of chemokine and cytokine messages and increased BALF levels of protein and 

PMNs. 

Fanucchi et al. (1998) and Wagner et al. (2001a,b) examined the synergistic effect of 

coexposure to O3 and endotoxin on the nasal transitional epithelium of rats that also was 

mediated, in part, by neutrophils. Fisher 344 rats intranasally instilled with endotoxin and 

exposed to 0.5 ppm O3, 8 h per day for 3 days developed mucous cell metaplasia in the nasal 

transitional epithelium, an area normally devoid of mucous cells; whereas, intratracheal 

instillation of endotoxin (20 μg) caused mucous cell metaplasia rapidly in the respiratory 

epithelium of the conducting airways. A synergistic increase of intraepithelial mucosubstances 

and morphological evidence of mucous cell metaplasia were found in rat maxilloturbinates upon 

exposure to both O3 and endotoxin, compared to each pollutant alone. A similar response was 

reported in OVA-sensitized Brown Norway rats exposed to 0.5 ppm O3, 8 h/day for 3 days 

(Wagner et al., 2002), indicating that coexposure to O3 and inflammatory biogenic substances 

like allergens (e.g., OVA) or bacterial endotoxin can augment epithelial and inflammatory 

responses in rat nasal passages. 

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In follow-up studies, Wagner et al. (2003) reported that coexposure of rats to O3 and 

endotoxin also enhanced epithelial and neutrophilic inflammatory responses in the pulmonary 

airways. Fisher 344 rats were intranasally instilled with endotoxin and exposed to 1.0 ppm O3 

for 8 h, which was repeated 24 h later. Three days after the last exposure, BALF was analyzed 

for inflammatory cells and secreted mucosubstances (mucin 5AC), and lung tissue was 

processed for morphometric analysis. Endotoxin instillation alone caused a dose-dependent 

increase in BALF neutrophils that was further increased 2-fold in O3-exposed rats given 20 μg 

endotoxin, the highest dose. Mucin glycoprotein 5AC also was increased in the BALF at this 

dose but not at lower endotoxin doses. Ozone exposure alone did not cause mucus 

hypersecretion, but it did potentiate mucus secretion in rats given both 2 and 20 μg endotoxin 

and increased intraepithelial mucosubstances 2-fold, which was further substantiated by 

significant increases in mucin gene (rMuc5AC) mRNA levels in the conducting airways. 

The effect of O3 modifying the biological potency of PM (diesel PM and carbon black) was 

examined by Madden et al. (2000) in rats. Reaction of National Institute of Standards and 

Technology (NIST) Standard Reference Material # 2975 diesel PM with 0.1 ppm O3 for 48 h 

increased the potency (compared to unexposed or air-exposed diesel PM) to induce neutrophil 

influx, total protein, and LDH in lung lavage fluid in response to intratracheal instillation. 

Exposure of the diesel PM to high, nonambient O3 concentration (1.0 ppm) attenuated the 

increased potency, suggesting destruction of the bioactive reaction products. Unlike the diesel 

particles, carbon black particles exposed to 0.1 ppm O3 did not exhibit an increase in biological 

potency, which suggested that the reaction of organic components of the diesel PM with O3 were 

responsible for the increased potency. 

Ulrich et al. (2002) investigated the effect of ambient PM from Ottawa Canada (EHC-93) 

on O3-induced inflammation. Male Wistar rats were exposed to 0.8 ppm O3 for 8 h and allowed 

to recover before intratracheal instillation of 0.5, 1.5, and 5 mg of EHC-93 in 0.3 mL of saline. 

The high concentrations of PM used were sufficient to induce pulmonary inflammation, which 

was not exacerbated by preexposure to O3. Rats from the combined exposure group had higher 

and more persistent lung lavage protein and albumin levels, as well as increased plasma 

fibrinogen levels, when compared to PM exposure alone. 

The interaction of PM and O3 was further examined in a murine model of OVA-induced 

asthma. Kobzik et al. (2001) investigated whether coexposure to inhaled, concentrated ambient 

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particles (CAPs) from Boston, MA and to O3 could exacerbate asthma-like symptoms. On days 

7 and 14 of life, half of the BALB/c mice used in this study were sensitized by i.p. injection of 

OVA and then exposed to OVA aerosol on three successive days to create the asthma phenotype. 

The other half received the i.p. OVA but were exposed to a phosphate-buffered saline aerosol 

(controls). The mice were further subdivided (n $ 61/group) and exposed for 5 h to CAPs, 

ranging from 63 to 1,569 μg/m3

, 0.3 ppm O3, CAPs + O3, or to FA. Pulmonary resistance and 

airway responsiveness to an aerosolized MCh challenge were measured after exposures. 

A small, statistically significant increase in RL and airway responsiveness, respectively, was 

found in both normal and asthmatic mice immediately after exposure to CAPs alone and to 

CAPs + O3 but not to O3 alone or to FA. By 24 h after exposure, the responses returned to 

baseline levels. There were no significant increases in airway inflammation after any of the 

pollutant exposures. In this well-designed study of a small-animal model of asthma, O3 and 

CAPs did not appear to be synergistic. In further analysis of the data using specific elemental 

groupings of the CAPs, the acutely increased RL was found to be associated with the AlSi 

fraction of PM. Thus, some components of concentrated fine particulate matter (PM2.5) may 

affect airway caliber in sensitized animals, but the results are difficult to extrapolate to people 

with asthma. 

Animal studies have examined the adverse cardiopulmonary effects of complex mixtures in 

urban and rural environments of Italy (Gulisano et al., 1997), Spain (Lorz and López, 1997), and 

Mexico (Vanda et al., 1998; Moss et al., 2001). Some of these studies have taken advantage of 

the differences in pollutant mixtures of urban and rural environments to report primarily 

morphological changes in the nasopharynx and lower respiratory tract (Gulisano et al., 1997; 

Lorz and López, 1997) of lambs and pigeons, respectively, after natural, continuous exposures to 

ambient pollution. Each study has provided evidence that animals living in urban air pollutants 

have greater pulmonary changes than those that would occur in a rural, and presumably cleaner, 

environment. However, these studies either did not report ambient O3 levels or reported only 

annual means. 

The study by Moss et al. (2001) examined the nasal and lung tissue of rats exposed 

(23 h/day) to Mexico City air for up to 7 weeks and compared them to controls similarly exposed 

to FA. No inflammatory or epithelial lesions were found using quantitative morphological 

techniques; however, the concentrations of pollutants were low. Extrapolation of these results to 

5-75
humans is restricted, however, by uncontrolled exposure conditions, small sample sizes, and 

other unknown exposure and nutritional factors in the studies in mammals and birds, and the 

negative studies in rodents. They also bring up the issue of which species of “sentinel” animals 

is more useful for predicting urban pollutant effects in humans. Thus, in these field studies, it is 

difficult to assign a specific role to any specific component of the mixture for the significant 

cardiopulmonary effects reported. 

Similar morphological changes (Calderón-Garcidueñas et al., 2000a, 2001) and chest X-ray 

evidence of mild lung hyperinflation (Calderón-Garcidueñas et al., 2000b) have been reported in 

children residing in urban and rural areas of Mexico City. The ambient air in urban areas, 

particularly in southwestern Mexico City, is a complex mixture of particles and gases, including 

high concentrations of O3 and aldehydes that previously have been shown to cause airway 

inflammation and epithelial lesions in humans (e.g., Calderón-Garcidueñas et al., 1992, 1994, 

1996) and laboratory animals (Morgan et al., 1986; Heck et al., 1990; Harkema et al., 1994, 

1997a,b). The described effects demonstrate a persistent, ongoing upper and lower airway 

inflammatory process and chest X-ray abnormalities in children residing predominantly in highly 

polluted areas. Again, extrapolation of these results to urban populations of the United States is 

difficult, because of the unique complex mixture of urban air in Mexico City, uncontrolled 

exposure conditions, and other unknown exposure and nutritional factors. 

5.4.4 Summary and Conclusions—Interactions of Ozone with Other 

Co-Occurring Pollutants 

It is difficult to summarize the role that O3 plays in exposure responses to binary mixtures, 

and even harder to determine its role in responses to multicomponent, complex atmospheres. 

Though the specific mechanisms of action of the individual pollutants within a mixture may be 

known, the exact bases for toxic interactions have not been elucidated clearly. Certain generic 

mechanisms that may underlie pollutant interactions: (1) physical, involving adsorption of one 

pollutant onto another and subsequent transport to more or less sensitive sites or to sites where 

one of the components of the mixture normally would not deposit in concentrated amounts 

(probably not involved in O3-related interactions); (2) production of secondary products that may 

be more toxicologically active than the primary materials, demonstrated or suggested in a 

number of studies as a basis for interaction between O3 and NO2 and between O3 and PM; 

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(3) biological or chemical alterations at target sites that affect response to O3 or the copollutant, 

H2SO4, which has been suggested to underlie interactions with mixtures of O3 and acid sulfates; 

(4) O3- or copollutant-induced physiological change, such as alteration in ventilation pattern, 

resulting in changes in the penetration or deposition of one pollutant when another is present. 

This has been implicated in enhanced responses to various O3-containing mixtures with exercise. 

Evaluation of interactions between O3 and copollutants is a complex task. Responses are 

dependent on a number of host and environmental factors, such that different studies using the 

same copollutants may show different types or magnitudes of interactions. The occurrence and 

nature of any interaction is dependent on the endpoint being examined and is also highly related 

to the specific conditions of each study, such as animal species, health status, exposure method, 

dose, exposure sequence, and the physicochemical characteristics of the copollutants. Because 

of this, it is difficult to compare studies, even those examining similar endpoints, that were 

performed under different exposure conditions. Thus, any description of interactions is really 

valid only for the specific conditions of the study in question and cannot be generalized to all 

conditions of exposure to a particular chemical mixture. Furthermore, it is generally not possible 

to extrapolate the effect of pollutant mixtures from studies of the effects of each component 

when given separately. In any case, what can be concluded from the database is that interactions 

of O3-containing mixtures are generally synergistic (antagonism has been noted in a few studies), 

depending on the various factors noted above and that O3 may produce more significant 

biological responses as a component of a mixture than when inhaled alone. Furthermore, 

although most studies have shown that interaction occurs only at higher than ambient 

concentrations with acute exposure, some have demonstrated interaction at more 

environmentally relevant levels (e.g., 0.05 to 0.1 ppm O3 with NO2) and with repeated exposures. 

5.5 EFFECTS OF OTHER PHOTOCHEMICAL OXIDANTS 

Peroxyacetyl nitrate and peroxypropionyl nitrate (PPN) are the most abundant non-O3 

oxidants in ambient air of industrialized areas, other than the inorganic nitrogenous oxidants 

such as NO2, and possibly HNO3. Ambient levels of PAN and PPN were reported to be 

decreasing over the 1990s, and available air quality data (Grosjean et al., 2001; Grosjean, 2003; 

Jakobi and Fabian, 1997) indicate that present peak concentrations of PAN and PPN in ambient 

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air from urban areas are in the low parts per billion range (e.g., <1 to 10 ppb). The levels found 

in nonurban areas are considerably lower (Gaffney et al., 1993). 

Reactions occur in the troposphere between O3 and hydrocarbons (e.g., D-limonene) to 

produce epoxides, hydroperoxides, and peroxides. The majority of the measured ambient 

hydroperoxides produced is H2O2, although a small amount of organic hydroperoxides (ROOH) 

also may be formed. Friedlander and Yeh (1998) have estimated that atmospheric aerosols can 

carry as high as 1 mM of H2O2 and ROOH (e.g., hydroxymethylhydroperoxide) in the associated 

water. In vitro cell and tissue damage are induced by high concentrations of liquid phase H2O2 

(50 μM to 1 mM). Morio et al. (2001) (see Table AX5-19) demonstrated that a 2 h exposure of 

10 and 20 ppb of inhaled H2O2 vapor can penetrate into lower lung regions where it causes 

inflammation. It is likely that hygroscopic components of PM transport ambient H2O2 into the 

lower lung and induce tissue injury as well. Exposure of rats to a H2O2-fine particle mixture 

(215 or 429 μg/m3

 (NH4)SO4) resulted in increased neutrophil influx, and production of 

inflammatory mediators by AMs (Morio et al., 2001). Hygroscopic secondary organic aerosols 

generated by O3/hydrocarbon reactions and their co-occurrence with H2O2 also provide another 

possible mechanism, yet to be validated, whereby H2O2 can be transported into the lower 

respiratory tract (e.g., Friedlander and Yeh, 1998). Interaction of inhaled O3 with unsaturated 

fatty acids on cell membranes and mucus in the airways generates epoxides, hydroperoxides, and 

secondary ozonation products such as HNE (see Section 5.2.1) 

Inhalation toxicological information on the effects of the non-O3 oxidants has been limited 

to a few studies on PAN, but at concentrations much higher (approximately 100- to 1,000-fold) 

than levels typically found in ambient air. Such acute high levels can cause changes in lung 

morphology, behavioral modifications, weight loss, and susceptibility to pulmonary infections. 

Therefore, acute toxicity of PAN is much lower than O3, and it is unlikely that present ambient 

PAN levels would affect pulmonary function responses to O3 (reviewed in Vyskocil et al., 1998). 

Cytogenetic studies indicate that PAN is not a potent mutagen, clastogen, or DNA damaging 

agent in mammalian cells in vivo or in vitro at concentrations several orders of magnitude higher 

than those generally encountered in ambient air in most cities (Vyskocil et al., 1998; Kligerman 

et al., 1995; Heddle et al., 1993). Some studies suggest that PAN may be a weak bacterial 

mutagen at concentrations much higher than exist in present urban atmospheres (DeMarini et al., 

2000; Kleindienst et al., 1990). 

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An additional level of complexity exists due to the possibility that other ambient oxidants 

may contribute to effects attributed to O3. As discussed in Chapter 2, both short-lived radicals 

and secondary particles containing highly polar compounds are generated in the troposphere by 

the same photochemical mechanisms that produce O3. It is plausible that, in addition to the 

direct effects of O3, health effects are produced by ambient exposures to these gaseous and 

particulate secondary compounds. Little is known regarding the composition of these reaction 

products, and little research has been undertaken evaluating their toxicologic effects. Due to the 

many oxidizing species present in the atmosphere, interpretation of toxicology data based on O3 

exposures alone have the potential for underestimating health effects of ambient oxidant 

mixtures. 

5.5.1 Summary and Conclusions—Effects of Other Photochemical Oxidants 

Concentrations of PAN and PPN (<1 to 10 ppb) in ambient air are unlikely to affect 

pulmonary function or cause DNA damage. Levels of 10-20 ppm H2O2 can penetrate to the 

lower lung directly or be transported there by PM, where inflammation can result; however, 

ambient H2O2 levels are typically < ~5 ppb. As toxicology studies of other photochemical 

oxidants are rare, quantitative scientific evaluations of possible health effects of environmental 

exposures cannot be completed at this time. 

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6. CONTROLLED HUMAN EXPOSURE STUDIES

OF OZONE AND RELATED

PHOTOCHEMICAL OXIDANTS

6.1 INTRODUCTION 

In the previous chapter, results of ozone (O3) studies in laboratory animals and in vitro test 

systems were presented. The extrapolation of results from animal studies is one mechanism by 

which information on potential adverse human health effects from exposure to O3 is obtained. 

More direct evidence of human health effects due to O3 exposure can be obtained through 

controlled human exposure studies of volunteers or through field and epidemiologic studies of 

populations exposed to ambient O3 (see Chapter 7). Controlled human exposure studies 

typically use fixed concentrations of O3 under carefully regulated environmental conditions and 

subject activity levels. This chapter discusses studies in which volunteers were exposed for up 

to 8 h to O3 concentrations ranging from 0.04 to 0.75 ppm O3 while at rest or during varying 

intensities of exercise. 

The majority of controlled human studies have investigated the effects of exposure to O3 in 

young nonsmoking healthy adults (18 to 35 years of age) performing continuous exercise (CE) 

or intermittent exercise (IE). Various combinations of O3 concentration, exercise routine, and 

exposure duration have been used in these studies. The responses to ambient O3 concentrations 

include decreased inspiratory capacity; mild bronchoconstriction; rapid, shallow breathing 

patterns during exercise; and symptoms of cough and pain on deep inspiration. Reflex inhibition 

of inspiration results in a decrease in forced vital capacity (FVC) and total lung capacity (TLC) 

and, in combination with mild bronchoconstriction, contributes to a decrease in the forced 

expiratory volume in 1 s (FEV1). In addition to physiological pulmonary responses and 

respiratory symptoms, O3 has been shown to result in airway hyperresponsiveness, epithelial 

permeability, and inflammation. 

The most salient observations from studies reviewed in the 1996 EPA Ozone Air Quality 

Criteria Document or O3 AQCD (U.S. Environmental Protection Agency, 1996) were that: 

(1) young healthy adults exposed to O3 concentrations $0.08 ppm develop significant reversible, 

6-1
transient decrements in pulmonary function if minute ventilation ( V0 E ) or duration of exposure is 

increased sufficiently, (2) children experience similar spirometric responses but lesser symptoms 

from O3 exposure relative to young adults, (3) O3-induced spirometric responses are decreased in 

the elderly relative to young adults, (4) there is a large degree of intersubject variability in 

physiologic and symptomatic responses to O3 but responses tend to be reproducible within a 

given individual over a period of several months, (5) subjects exposed repeatedly to O3 for 

several days develop a tolerance to successive exposures, as demonstrated by an attenuation of 

responses that is lost after about a week without exposure, and (6) acute O3 exposure initiates an 

inflammatory response which may persist for at least 18 to 24 h postexposure. 

There are several important limitations associated with these clinical studies: (1) the 

ability to study only short-term, acute effects; (2) difficulties in trying to link short-term effects 

with long-term consequences; (3) the use of a small number of volunteers that may not be 

representative of the general population; and (4) statistical limitations associated with the small 

sample size. Sample size affects the power of a study, and having a small number of samples 

causes a risk of Type II error, i.e., the incorrect conclusion that no difference exists between 

treatments or groups when comparisons are not significantly different. This affects the 

confidence in estimates of a minimum O3 concentration at which some degree of pulmonary 

impairment will occur in both the general population and susceptible subpopulations. As a 

result, the conclusions drawn from many of the studies cited in this chapter may underestimate 

the presence of responses at low O3 concentrations and low activity levels. 

Most of the scientific information summarized in this chapter comes from the literature 

published since the 1996 O3 AQCD (U.S. Environmental Protection Agency, 1996). In addition 

to further study of physiological pulmonary responses and symptoms of breathing discomfort, 

much of this literature has focused on mechanisms of inflammation and cellular responses to 

injury induced by O3 inhalation. A more thorough discussion and review of this literature 

appears in Annex AX6 of this document. In summarizing the literature, effects are described if 

they are statistically significant at a probability (p-value) of less than 0.05; otherwise, trends are 

noted as such. 

As spirometry typically improves in healthy young adults with exercise exposures to 

filtered air (FA), the term “O3-induced” is used herein and in the annex to designate effects that 

have been corrected for responses during FA exposures. For healthy adults, an O3-induced 

6-2
change in lung function is the difference between the decrement experienced with O3 exposure 

and the improvement observed with FA exposure. However, the distinction between an O3­

induced change and a post- versus preexposure change is particularly important in individuals 

with respiratory disease who may experience exercise-induced decrements in pulmonary 

function during both FA and O3 exposures. Hence, in subjects with respiratory disease, exerciseinduced responses could be mistaken for O3-induced responses in the absence of a correction for 

FA responses. 

6.2 PULMONARY FUNCTION EFFECTS OF OZONE EXPOSURE 

IN HEALTHY SUBJECTS 

6.2.1 Introduction 

As reviewed in the 1986 and 1996 O3 AQCD’s (U.S. Environmental Protection Agency, 

1986, 1996), 0.5 ppm is the lowest O3 concentration at which statistically significant reductions 

in FVC and FEV1 have been reported in sedentary subjects. On average, young adults (n = 23; 

mean age, 22 yrs) exposed at rest for 2 h to 0.5 ppm O3 had O3-induced decrements of ~4% in 

FVC and ~7% in FEV1 (Folinsbee et al., 1978; Horvath et al., 1979). During exercise, 

spirometric and symptoms responses are observed at lower O3 concentrations. For acute 

exposures of 1 to 2 h to /0.12 ppm O3, if V0 E is sufficiently increased by exercise, healthy human 

subjects generally experience (a) decreases in TLC, inspiratory capacity (IC), FVC, FEV1, mean 

forced expiratory flow from 25% to 75% of FVC (FEF25-75), and tidal volume (VT) and 

(b) increases in specific airways resistance (sRaw), breathing frequency (fB), and airway 

responsiveness. These exposures also cause symptoms of cough; pain on deep inspiration; rapid, 

shallow breathing patterns during exercise; throat irritation; and wheezing. With exposures of 4­

to 8-h in duration, statistically significant pulmonary function and symptoms responses are 

observed at lower O3 concentrations and lower V0 E than in shorter duration studies. 

A large body of data regarding the interdependent effects of O3 concentration (C), minute 

ventilation ( V0 E ), and duration of exposure (time, T) on pulmonary responses was assessed in the 

1986 and 1996 O3 AQCD’s (U.S. Environmental Protection Agency, 1986, 1996). In an attempt 

to describe O3 dose-response characteristics, acute responses were modeled as a function of total 

6-3
inhaled O3 dose (C × T × V0 E ), which was found to be a better predictor of response than O3 

concentration, V0 E, or T of exposure, alone, or as a combination of any two of these factors. For 

example, in an analysis of 6 studies with 1- to 2-h exposures to between 0.12 and 0.18 ppm O3 

with exercise, Folinsbee et al. (1988) found a good correlation (r = 0.81) between total inhaled 

O3 dose and FEV1 decrements. Other studies, however, reported that O3 concentration appears to 

be more important than either V0 E or T in determining pulmonary responses during acute 

exposures at constant concentration (Adams et al., 1981, 1987; Folinsbee et al, 1978; Hazucha, 

1987; Larsen et al., 1991). When considering variable O3 concentrations, where response is not 

a linear function of dose, the use of O3 dose as a predictor of response becomes more complex, 

if appropriate at all (e.g., see Section 6.2.4). 

6.2.2 Acute Exposure for Up to 2 h 

With heavy CE ( V0 E = 89 L/min), an O3-induced decrement of 9.7% in FEV1 has 

been reported for healthy young adults (n = 17; age, 24 ± 3 yrs) exposed for only 1 h to 

0.12 ppm O3 (Gong et al., 1986). With moderate-to-heavy IE (15 min intervals of rest and 

exercise [ V0 E = 68 L/min]), McDonnell et al. (1983) reported a physiologically small, but 

significant, O3-induced decrement of 3.4% in FEV1 for young healthy adults (n = 22, age, 

22 ± 3 yrs) exposed for 2 h to 0.12 ppm O3. Using the same 2 h IE exposure protocol, Linn et al. 

(1986) found no statistically significant spirometic responses at O3 concentrations of 0.16 ppm 

and lower. However, the subjects in the Linn et al. (1986) study were potentially exposed 

concurrently in Los Angeles to ambient O3 levels of between 0.12 and 0.16 ppm and were on 

average 3 yrs older than the subjects in the McDonnell et al. (1983) study. (The attenuating 

effects of increasing age and repeated O3 exposures are discussed in Sections 6.5.1 and 6.6, 

respectively.) The disparities between the Linn et al. (1986) and McDonnell et al. (1983) studies 

demonstrate the difficulty in determining a no-effect-level for O3 based on relatively small study 

populations. 

Studies analyzing large data sets ($300 subjects) provide better predictive ability of acute 

changes in FEV1 at low levels of O3 and V0 E than is possible via comparisons between smaller 

studies. Such an analysis was performed by McDonnell et al. (1997), who examined FEV1 

responses in 485 healthy white males (18 to 36 years of age; subjects recruited from the area 

6-4
around Chapel Hill, NC) exposed once for 2 h to O3 concentrations of up to 0.40 ppm at rest or 

with IE. Decrements in FEV1 were modeled by sigmoid-shaped curve as a function of subject 

age, O3 concentration, V0 E , and duration of exposure. Regarding applicability to the general 

population, the McDonnell et al. (1997) model has an apparent limitation of considering only 

data for young adult white males. However, two other analyses (93 black females; 94 white 

females; 92 black males; 93 white males, a subset of subjects included in the McDonnell et al. 

study) found no significant gender or race effects on spirometric responses to O3 exposure (Seal 

et al., 1993, 1996). It should be emphasized that the McDonnell et al. (1997) model only 

predicts average responses. For exposure conditions causing a predicted average FEV1 

decrement of 10%, individual decrements range from approximately 0 to 40% (see Figure 1 of 

McDonnell et al., 1997). (The reader is referred to Sections 6.5.1 and 6.5.4 for further 

discussion of this model and its prediction of age and physical activity effects on FEV1 

responses.) 

In a more recent study, McDonnell et al. (1999) also reported a model predicting average 

symptom responses from O3 exposure. Unfortunately, neither of these papers (McDonnell et al., 

1997, 1999) provide predictions of intersubject variability in response. Ultman et al. (2004) 

recently reported pulmonary responses in 60 young heathy nonsmoking adults (32 M, 28 F) 

exposed to 0.25 ppm O3 for 1 h with CE at a target V0 E of 30 L/min. Consistent with findings 

reported in the 1996 O3 criteria document, considerable intersubject variability in FEV1 

decrements was reported by Ultman et al. (2004) with responses ranging from a 4% 

improvement to a 56% decrement. One-third of the subjects had FEV1 decrements of >15% and 

7% of the subjects had decrements of >40%. (Section 6.4 of this Chapter discusses intersubject 

variability in response to O3 exposure.) 

In addition to overt effects of O3 exposure on the large airways indicated by spirometric 

responses, O3 exposure also affects the function of the small airways and parenchymal lung. 

Foster et al. (1993, 1997) examined the effect of O3 on ventilation distribution in healthy adult 

males. In healthy nonsmoking males (26.7 ± 7 years old) exposed to FA or 0.33 ppm O3 for 2 h 

with IE, there was a significant reduction in ventilation to the lower lung (31% of lung volume) 

and significant increases in ventilation to the upper- and middle-lung regions relative to the FA 

values in 7 of the 9 subjects (Foster et al., 1993). In another study, 15 healthy nonsmoking 

6-5
males (25.4 ± 2 years old) were exposed to FA or 0.35 ppm O3 for 2.2 h with IE (Foster et al., 

1997). Following O3 exposure, an inert gas washout was delayed and resembled a twocompartment washout, whereas pre-O3 exposure a log-linear gas clearance as a function of 

expired volume resembled a single-compartment washout. The pronounced slow phase of gas 

washout occurring post-O3, represented a 24% decrease in the washout rate relative to pre-O3. 

At 24-h post-O3, 6 of the 12 subjects still had (or developed) a delayed washout relative to the 

pre-O3 maneuver. This suggests a prolonged O3 effect on the small airways and ventilation 

distribution in some individuals. 

6.2.3 Prolonged Ozone Exposures 

In the exposure range of 0.08 to 0.16 ppm O3, a number of studies using moderate 

quasi continuous exercise (QCE; 50 min exercise and 10 min rest per h) for 4 to 8 h have 

shown significant responses under the following conditions: 0.16 ppm for 4 h with QCE 

at V0 E . 40 L/min (Folinsbee et al., 1994), 0.08 to 0.12 ppm for 6.6 h with QCE at V0 E . 35 to 

40 L/min (Adams, 2002; Adams, 2003a; Folinsbee et al., 1988; Horstman et al., 1990), and 

0.12 ppm for 8 h of IE (30 min per h) at V0 E . 40 L/min (Hazucha et al., 1992). Symptoms and 

spirometric responses increased with duration of exposure, O3 concentration, and total V0 E . 

Airway resistance is only modestly affected with moderate or even heavy exercise combined 

with O3 exposure (Folinsbee et al., 1978; McDonnell et al., 1983; Seal et al., 1993). 

6.2.3.1 Effect of Exercise Ventilation Rate on FEV1 Response to 6.6 h Ozone Exposure 

It is well established that response to O3 exposure is a function of V0 E in studies of 2 h or 

less in duration (See Section AX6.2.2). It is reasonable to expect that response to a prolonged 

6.6-h O3 exposure is also a function of V0 E, although quantitative analyses are lacking. Data 

from five similar prolonged exposure studies are available for evaluation of FEV1 responses as a 

function of exercise V0 E (Adams, 2000; Adams and Ollison, 1997; Folinsbee et al., 1988, 1994; 

Horstman et al., 1990). Each of these studies exposed similarly aged subjects (mean ages 22 to 

25 yrs) to 0.12 ppm O3 for 6.6 h. In total, ten sets of mean FEV1 decrements were available for 

exercise 0 VE VE ranging from 20 to 43 L/min, although no data were available for 0 between

20 and 30 L/min (data illustrated in Figure AX6-2). As in 2-h exposure studies, FEV1 

6-6
0

decrements are a function of V0 E in prolonged 6.6-h exposure studies as indicated by a significant 

correlation between these variables (Pearson, r = 0.95, p < 0.001; Spearman, r = 0.84, p < 0.01). 

6.2.3.2 Exercise Ventilation Rate as a Function of Body/Lung Size on FEV1 Response 

to 6.6 h Ozone Exposure 

Based on the assumption that the total inhaled O3 dose (product of O3 concentration, 

exposure duration, and 0 VE VE ) is proportional to the lung size, exercise 0 is typically selected to be

a multiple of body surface area (BSA) or FVC. Data from several recent studies do not support 

the contention that V0 E should be normalized. In an analysis of data from 485 young adults, 

McDonnell et al. (1997) found that any effect of BSA, height, or baseline FVC on percent 

decrement in FEV1 was small to nonexistent. This is consistent with Messineo and Adams 

(1990), who compared pulmonary function responses in young adult women having small 

(n = 14) or large (n = 14) lung sizes (mean FVC of 3.74 and 5.11 L, respectively) and found no 

significant group difference in FEV1 decrements. For 30 subjects (15M, 15F) exposed to 

0.12 ppm O3 for 6.6 h, Adams (2000) also reported that FEV1 responses were more closely 

VE related to 0 than to V0 E normalized to BSA. The O3 dosimetry study of Bush et al. (1996) 

suggested that normalization of the O3 dose might more appropriately be a function of anatomic 

dead space. Ozone penetrates deeper into the lungs of individuals with larger conducting airway 

volumes, however, differences in FEV1 responses between subjects exposed for 2 h to 0.25 ppm 

O3 (VE = 30 L/min) do not appear to be explained by intersubject differences in the fraction of 

inhaled O3 retained in the lung (Ultman et al., 2004). 

6.2.3.3 Comparison of 2 h IE to 6.6 h O3 Exposure Effects on Pulmonary Function 

Adams (2003b) examined whether prolonged 6.6-h QCE exposure to a relatively low O3 

concentration (0.08 ppm) and a 2-h IE exposure at a relatively high O3 concentration (0.30 ppm) 

elicited consistent individual subject FEV1 responses. Individual subject O3 exposure 

reproducibility was first examined via a regression plot of the postexposure FEV1 response to the 

6.6-h chamber exposure as a function of postexposure FEV1 response to the 2-h IE chamber 

exposure. The resulting R2

 of 0.40, although statistically significant, was substantially less than 

that observed in a comparison of individual FEV1 response to the two 2-h IE exposures by 

6-7
chamber and face mask, respectively (R2

 = 0.83). The Spearman rank order correlation for the 

chamber 6.6-h and chamber 2-h exposure comparison was also substantially less (0.49) than that 

obtained for the two 2-h IE exposures (0.85). The primary reason for the greater variability in 

the chamber 6.6-h exposure FEV1 response as a function of that observed for the two 2-h IE 

exposures is very likely related to the increased variability in response upon repeated exposure 

to O3 concentrations lower than 0.18 ppm (R = 0.57, compared to a mean R of 0.82 at higher 

concentrations) reported by McDonnell et al. (1985). This rationale is supported by the lower r2 

(0.40) observed by Adams (2003b) for the FEV1 responses found in 6.6 h chamber and face 

mask exposures to 0.08 ppm O3, compared to an r2

 of 0.83 observed for responses found at 0.30 

ppm O3. 

6.2.4 Triangular Ozone Exposures 

To further explore the factors that determine responsiveness to O3, Hazucha et al. (1992) 

designed a protocol to examine the effect of varying, rather than constant, O3 concentrations. 

Subjects were exposed to an O3 level that increased linearly from 0 to 0.24 ppm for the first 4 h 

and then decreased linearly from 0.24 to 0 ppm over the second 4 h of the 8 h exposure 

(triangular concentration profile) and to a constant level exposure of 0.12 ppm O3 for 8 h. While 

total inhaled O3 doses for the constant and the triangular concentration profile were almost 

identical, the FEV1 response was dissimilar. For the constant 0.12 ppm O3 exposure, FEV1 

declined ~5% by the fifth hour and then remained at that level. With the triangular O3 

concentration profile, there was minimal FEV1 response over the first 3 h followed by a rapid 

decrease in FEV1 (!10.3%) over the next 3 h. During the seventh and eighth hours, mean FEV1 

decrements improved to !6.3% as the O3 concentration decreased from 0.12 to 0.00 ppm 

(mean = 0.06 ppm). (The reader is referred to Figure AX6-3 for illustration of FEV1 responses 

during square-wave and triangular O3 exposures). 

Adams (2003a) used a less abrupt triangular O3 exposure profile (ranging from 0.03 to 

0.15 ppm) having an average exposure concentration of only 0.08 ppm, assumed to be typical of 

outdoor ambient conditions. Postexposure values for FEV1 and symptoms were not significantly 

different between the 6.6 h triangular and a square-wave 0.08 ppm O3 exposure. During the 

triangular exposure, however, FEV1 responses became statistically significant after 4.6 h, 

6-8
whereas they were not significant until 6.6 h during the square wave exposure (Adams, 2003a). 

Perhaps due to the lower O3 concentrations, evidence of FEV1 response recovery with the 

triangular exposure was less pronounced than as observed by Hazucha et al. (1992). Figure 6-1 

illustrates the average O3-induced FEV1 responses and the O3 exposure schemes for the Adams 

(2003a) and Hazucha et al. (1992) studies. 

Figure 6-1. Ozone-induced changes in FEV1 (top panel) and O3 concentration profiles 

(bottom panel) as a function of exposure duration. Open (") and closed (!) 

circles illustrate average data from Hazucha et al. (1992) and Adams (2003a), 

respectively. For clarification, the “O3-induced changes in FEV1” are the FEV1 

responses following O3 exposure minus the FEV1 responses following FA. 

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In a very recent chamber study, Adams (2006) employed the earlier design of square-wave 

and triangular exposure profiles. The O3 concentrations for the 6.6-h exposures were 0.00, 0.06, 

and 0.08 ppm for square-wave profiles and averaged 0.04, 0.06, and 0.08 ppm for triangular 

profiles. The 6.6-h postexposure responses from the 0.08 ppm O3 (average) triangular exposure 

did not differ significantly from those observed in the 0.08 ppm O3 square-wave exposure. 

However, FEV1 and symptoms were significantly different from preexposure at 4.6 h (when 

the O3 concentration was 0.15 ppm) in the triangular exposure, but not until 6.6 h in the squarewave exposure. These observations have confirmed and expanded findings of the earlier study 

(Adams, 2003a) showing a clear divergence in the peak and hourly responses in FEV1 between 

square-wave and triangular exposure profiles at the 0.08 ppm level. However, at the lower O3 

concentration of 0.06 ppm, no temporal pattern differences in FEV1 responses between squarewave and triangular exposure profiles could be discerned. The author concluded that the results 

support previous evidence that O3 concentration has a greater singular effect in the total inhaled 

O3 dose than do V0 E and exposure duration. The author observed no significant differences in 

FEV1 and symptom responses compared to FA for the 0.04 and 0.06 ppm exposures. 

With square-wave O3 exposures between 0.08 to 0.12 ppm, FEV1 decrements may increase 

with time of exposure (and O3 dose) or reach a plateau (Horstman et al., 1990; McDonnell et al., 

1991). For the triangular exposures used by Hazucha et al. (1992) and Adams (2003a, 2006), 

maximal FEV1 responses occurred 1 h to 2 h after peak O3 concentration and 1 h to 2 h before 

the maximal O3 dose occurred (at the end of the O3 exposure). These three studies suggest that a 

triangular exposure profile can potentially lead to higher FEV1 responses than square-wave 

profiles at overall equivalent O3 doses. 

6.2.5 Mechanisms of Pulmonary Function Responses 

Inhalation of O3 for several hours while physically active elicits both subjective respiratory 

tract symptoms and acute pathophysiologic changes. The typical symptomatic responses 

consistently reported in studies are that of airway irritation, cough, and pain on deep inspiration. 

Depending on the individual’s responsiveness to O3, this can be accompanied by several 

pathophysiologic changes, e.g., decrements in lung capacities and volumes, bronchoconstriction, 

airway hyperresponsiveness, airway inflammation, immune system activation, and epithelial 

injury. The severity of symptoms and the magnitude of response depend on inhaled dose, O3 

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sensitivity of an individual, and the extent of tolerance resulting from the individual’s previous 

exposures. The development of effects is time-dependent during both exposure and recovery 

periods, with considerable overlap of evolving and receding effects. The time sequence, 

magnitude and the type of responses of this complex series of events, both in terms of 

development and recovery, indicate that several mechanisms, activated at different times of 

exposure, must contribute to the overall lung function response (U.S. Environmental Protection 

Agency, 1996). 

Available information on recovery from O3 exposure indicates that an initial phase of 

recovery proceeds relatively rapidly, and some 40 to 65% of the acute spirometric and symptom 

response appears to occur within about 2 h (Folinsbee and Hazucha, 1989). Following a 2-h 

exposure to 0.4 ppm O3 with IE, Nightingale et al. (2000) observed a 13.5% decrement in FEV1. 

By 3 h postexposure; however, only a 2.7% FEV1 decrement persisted, as illustrated in 

Figure 6-2. A similar postexposure recovery in FVC was also observed. Gerrity et al. (1993) 

suggested that, for healthy young adults, transient increases in mucus clearance (mediated by 

cholinergic receptors) due to O3 exposure may be coincident to pulmonary function responses, 

i.e., the transient increases in clearance and decrements in lung function return to baseline values 

within 2 to 3 h postexposure. However, there is some indication that the spirometric responses, 

especially at higher O3 concentrations, are not fully recovered within 24 h (Folinsbee and 

Horvath, 1986; Folinsbee et al., 1998). In hyperresponsive individuals, the recovery takes 

longer, as much as 48 h, to return to baseline values. Collectively, these observations suggest 

that there is a rapid recovery of O3-induced spirometric responses and symptoms, which may 

occur during resting exposure to O3 (Folinsbee et al., 1977) or as O3 concentration is reduced 

during exposure (Hazucha et al., 1992), and a slower phase, which may take at least 24 h to 

complete (Folinsbee and Hazucha, 2000). Repeated exposure studies at higher concentrations 

typically show that FEV1 response to O3 is enhanced on the second of several days of exposure 

(Table AX6-8). This enhanced response suggests a residual effect of the previous exposure, 

about 22 h earlier, even though the preexposure spirometry may be the same as on the previous 

day. The absence of the enhanced response with repeated exposure at lower O3 concentrations 

may be the result of a more complete recovery or less damage to pulmonary tissues (Folinsbee 

et al., 1994). 

6-11
Figure 6-2. Recovery of FEV1 responses following a 2 h exposure to 0.4 ppm O3 with IE. 

Note that the 2 h exposure period is indicated by the Pre to 0 h interval. 

Immediately postexposure, FEV1 was significantly (**p < 0.001) decreased. 

At 3 h postexposure, FEV1 was at 97% of the preexposure value. 

Adapted from Nightingale et al. (2000). 

6.2.5.1 Pathophysiologic Mechanisms 

Breathing pattern changes 

Human studies consistently report that inhalation of O3 alters the breathing pattern without 

significantly affecting minute ventilation. A progressive decrease in tidal volume and a 

“compensatory” increase in frequency of breathing to maintain steady minute ventilation during 

exposure suggests a direct modulation of ventilatory control. These changes parallel a response 

of many animal species exposed to O3 and other lower airway irritants (Tepper et al., 1990). 

Bronchial C-fibers and rapidly adapting receptors appear to be the primary vagal afferents 

responsible for O3-induced changes in ventilatory rate and depth in both humans (Folinsbee and 

Hazucha, 2000) and animals (Coleridge et al., 1993; Hazucha and Sant'Ambrogio, 1993; 

Schelegle et al., 1993). 

6-12
The potential modulation of breathing pattern by activation of sensory afferents located in 

extrathoracic airways by O3 has not yet been studied in humans. Nasal-only O3 exposure of rats 

produces changes in breathing pattern that are similar to changes observed in humans (Kleinman 

et al., 1999). 

Symptoms and lung function changes 

As discussed, in addition to changes in ventilatory control, O3 inhalation by humans can 

also induce a variety of symptoms, reduce vital capacity (VC) and related functional measures, 

and increase airway resistance. 

Schelegle et al. (2001) demonstrated that the reduction in VC due to O3 exposure is a reflex 

action and not a voluntary termination of inspiration as result of discomfort. They reported 

that O3-induced symptom responses (mediated in part by bronchial C-fibers) are substantially 

reduced by inhaled topical anesthetic. However, the anesthetic had a minor and irregular effect 

on pulmonary function decrements and tachypnea. Since respiratory symptom responses were 

largely abolished, these findings support reflex inhibition of VC due to stimulation of both 

bronchial and pulmonary C-fibers. 

The involvement of nociceptive bronchial C-fibers modulated by opioid receptors 

in limiting maximal inspiration and eliciting subjective symptoms in humans was studied 

by Passannante et al. (1998). Sufentanil (an opioid agonist and analgesic) rapidly 

reversed O3-induced symptom responses and reduced spirometric decrements in “strong” 

responders. The incomplete recovery in FEV1 following sufentanil administration, however, 

suggests involvement of non-opioid receptor modulated mechanisms as well. Interestingly, 

naloxone (an opioid receptor antagonist) had no significant effect on FEV1 decrements in “weak” 

responders. Plasma levels of $-endorphin (a potent pain suppressor) were not related to 

O3 responses. 

Airway hyperreactivity 

In addition to limitation of maximal inspiration and its effects on other spirometric 

endpoints, activation of airway sensory afferents also plays a role in receptor-mediated 

bronchoconstriction and an increase in airway resistance. Despite this common mechanism, 

post-O3 pulmonary function changes and either early or late airway hyperresponsiveness (AHR) 

6-13
to inhaled aerosolized methacholine or histamine are poorly correlated either in time or 

magnitude. Fentanyl and indomethacin, the drugs that have been shown to attenuate O3-induced 

lung function decrements in humans, did not prevent induction of AHR when administered to 

guinea pigs prior to O3 exposure (Yeadon et al., 1992). Neither does post-O3 AHR seem to be 

related to baseline airway responsiveness. These findings imply that the mechanisms are either 

not related or are activated independently in time. Animal studies (with limited support from 

human studies) have suggested that an early post-O3 AHR is, at least in part, vagally mediated 

(Freed, 1996) and that stimulation of C-fibers can lead to increased responsiveness of bronchial 

smooth muscle independently of systemic and inflammatory changes which may be even absent 

(Joad et al., 1996). An in vitro study of isolated human bronchi has shown that O3-induced 

airway sensitization involves changes in smooth muscle excitation-contraction coupling 

(Marthan, 1996). Characteristic O3-induced inflammatory airway neutrophilia, which at one 

time was considered a leading AHR mechanism, has been found in a murine model to be only 

coincidentally associated with AHR, i.e., there was no cause and effect relationship (Zhang et al., 

1995). However, this observation does not rule out possible involvement of other cells (such as 

eosinophils or T-helper cells) in AHR modulation. There is some evidence that release of 

inflammatory mediators by these cells can sustain AHR and bronchoconstriction. In vitro and 

animal studies have also suggested that airway neutral endopeptidase activity can be a strong 

modulator of AHR (Marthan et al., 1996; Yeadon et al., 1992). Late AHR observed in some 

studies is plausibly due to sustained damage of the airway epithelium and continual release of 

inflammatory mediators (Foster et al., 2000). Thus, O3-induced AHR appears to be a product of 

multiple mechanisms acting at different time periods and levels of the bronchial smooth muscle 

signaling pathways (effects of O3 on AHR are described in Section 6.8). 

6.2.5.2 Mechanisms at a Cellular and Molecular Level 

Stimulation of vagal afferents by O3 and reactive products, the primary mechanism of lung 

function impairment, is enhanced and sustained by what can be considered in this context to be 

secondary mechanisms activated at a cellular and molecular level. The complexity of these 

mechanisms is beyond the scope of this section, and the reader is directed to Section 6.9 of this 

chapter for more detail. A comprehensive review by Mudway and Kelly (2000) discusses the 

cellular and molecular mechanisms of O3-induced pulmonary response in great detail. 

6-14
Stimulation of bronchial C-fibers by O3 not only inhibits maximal inspiration but, through 

local axon reflexes, induces neurogenic inflammation. This pathophysiologic process is 

characterized by release of tachykinins and other proinflammatory neuropeptides. Ozone 

exposure has been shown to elevate the C-fiber-associated tachykinin, substance P, in human 

bronchial lavage fluid (Hazbun et al. 1993) and to deplete neuropeptides synthesized and 

released from C-fibers in human airway epithelium rich in substance P-immunoreactive axons. 

Substance P and other transmitters are known to induce granulocyte adhesion and subsequent 

transposition into the airways, increase vascular permeability and plasma protein extravasation, 

cause bronchoconstriction, and promote mucus secretion (Solway and Leff, 1991). Although the 

initial pathways of neurogenic, antigen-induced, and innate immune-mediated inflammation are 

not the same, they eventually converge, leading to further amplification of airway inflammatory 

processes by subsequent release of cytokines, eicosanoids, and other mediators. Significantly 

negative correlations between O3-induced leukotriene (LTC4/D4/E4) production and spirometric 

decrements (Hazucha et al., 1996) and an increased level of postexposure PGE2, a mediator 

known to stimulate bronchial C-fibers, show that these mediators play an important role in lung 

function impairment due to O3 exposure (Mohammed et al., 1993; Hazucha et al., 1996). The 

reported post O3 exposure dysfunction of small airways assessed by decrement in FEF25-75 

(Weinman et al., 1995; Frank et al., 2001) is likely induced by both neurogenic and 

inflammatory mediators, since the density of bronchial C-fibers is much lower in the small than 

large airways. Also, because of the relative slowness of inflammatory responses as compared to 

reflex effects, O3-triggered inflammatory mechanisms are unlikely to initially contribute to 

progressive lung function reduction. It is plausible, however, that when fully activated, they 

sustain and possibly further aggravate already impaired lung function. Indeed, a prolonged 

recovery of residual spirometric decrements following the initial rapid improvement after 

exposure termination could be due to slowly resolving airway inflammation. Bronchial 

biopsies performed 6 h postexposure have shown that O3 caused a significant decrease in 

immunoreactivity to substance P in the submucosa (Krishna et al., 1997). A strong negative 

correlation with FEV1 also suggests that the release of substance P may be an important 

contributing mechanism to persistent post-O3 bronchoconstriction (Krishna et al., 1997). 

Persistent spirometric changes observed for up to 48 h postexposure could plausibly be 

6-15
sustained by the inflammatory mediators, many of which have bronchoconstrictive properties 

(Blomberg et al., 1999). 

6.3 SUBJECTS WITH PREEXISTING DISEASE 

Individuals with respiratory disease are of primary concern in evaluating the health effects 

of O3 because even a small change in function is likely to have more impact on a person with 

reduced reserve, i.e., O3-induced effects are superimposed on preexisting pulmonary impairment. 

6.3.1 Subjects with Chronic Obstructive Pulmonary Disease 

For patients with COPD performing light to moderate IE, no decrements in pulmonary 

function were observed after 1- and 2-h exposures to #0.30 ppm O3 (Kehrl et al., 1985; Linn 

et al., 1982a, 1983a; Solic et al., 1982), and only small decreases in forced expiratory volume 

were observed for 3-h exposures of chronic bronchitics to 0.41 ppm O3 (Kulle et al., 1984). 

More recently, Gong et al. (1997a) found no significant difference in response between agematched controls and COPD patients to a 4 h exposure to 0.24 ppm O3 with IE. Although the 

clinical significance is uncertain, small transient decreases in arterial blood oxygen saturation 

have also been observed in some of these studies. 

6.3.2 Subjects with Asthma 

Based on studies reviewed in the 1996 O3 AQCD (U.S. Environmental Protection Agency, 

1996), asthmatic subjects appear to be at least as sensitive to acute effects of O3 as healthy 

nonasthmatic subjects. 

Several recent studies support a tendency for slightly increased spirometric responses in 

mild asthmatic versus healthy subjects. Alexis et al. (2000) reported reductions in FVC (12%, 

10%) and FEV1 (13%, 11%) for 13 mild asthmatic and 9 healthy subjects, respectively, exposed 

to 0.4 ppm O3 for 2 h with IE ( V0 E = 30 L/min). The FVC and FEV1 responses were attenuated 

by indomethacin in the healthy subjects, but not in the asthmatic subjects. As assessed by the 

magnitude of reductions in mid-flows (viz. FEF25, FEF50, FEF60p, FEF75) following O3 exposure, 

the small airways tended to be more affected in the asthmatic than in the healthy subjects. In a 

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larger study, Jörres et al. (1996) exposed 24 asthmatics, 12 allergic rhinitics, and 10 healthy 

subjects to 0.25 ppm O3 for 3 h with IE ( V0 E = 30 L/min). The O3-induced FEV1 decrements 

tended to be greater in the diseased populations (allergic rhinitis, 14.1%; asthmatic, 12.5%; 

healthy controls, 10.2%). Also, Scannell et al. (1996) exposed 18 asthmatics to 0.2 ppm O3 for 

4 h with IE ( V0 E . 25 L/min/m2 BSA). An O3-induced increase in sRaw tended to be greater in 

the asthmatics compared to 81 healthy subjects who underwent similar experimental protocols 

(Aris et al., 1995; Balmes et al., 1996). 

Similar O3-induced spirometric responses of asthmatic and healthy subjects are suggested 

by some studies. The Scannell et al. (1996) study of 18 asthmatics reported FEV1 and FVC 

decrements that were similar to 81 healthy subjects (Aris et al., 1995; Balmes et al., 1996). 

Similar group decrements in FEV1 and FVC were also reported by Hiltermann et al. (1995) for 

6 asthmatic and 6 healthy subjects exposed to 0.4 ppm O3 for 2 h with light IE. Basha et al. 

(1994) also reported similar spirometric responses between 5 asthmatic and 5 healthy subjects 

exposed to 0.2 ppm O3 for 6 h with IE. The lack of significant differences in the Hiltermann 

et al. (1995) and Basha et al. (1994) studies is not very compelling, however, given the 

extremely small sample sizes and corresponding lack of statistical power. The Basha et al. 

(1994) study was also confounded by the asthmatics having an average preexposure FEV1 

that was about 430 mL lower (a 12% difference) on the O3-exposure day relative to the 

FA-exposure day. 

One other study has reported results suggesting that some asthmatics may have smaller 

O3-induced FEV1 decrements relative to healthy subjects (3% versus 8%, respectively) when 

exposed to 0.2 ppm O3 for 2 h with IE (Mudway et al., 2001). However, the asthmatics in the 

Mudway et al. (2001) study also tended to be older than the healthy subjects, which could 

partially explain their lesser response. 

In a longer exposure duration (7.6 h) study, Horstman et al. (1995) exposed 17 mild-tomoderate asthmatics and 13 healthy controls to 0.16 ppm O3 or FA with quasi continuous 

exercise ( V0 E .30 L/min). The FEV1 decrement observed in the asthmatics was significantly 

larger than in the healthy subjects (19% versus 10%, respectively). There was also a notable 

tendency for a greater O3-induced decrease in FEF25-75 in asthmatics relative to the healthy 

subjects (24% versus 15%, respectively). A significant positive correlation in asthmatics was 

6-17
also reported between O3-induced spirometric responses and baseline lung function, i.e., 

responses increased with severity of disease. 

With repeated O3 exposures, asthmatic subjects, like healthy subjects (see Section 6.6), 

develop tolerance. Gong et al. (1997b) exposed 10 asthmatics to 0.4 ppm O3, 3 h per day with 

IE ( V0 E .32 L/min), for 5 consecutive days. Symptom and spirometric responses were greatest 

on the first (!35 % FEV1) and second (!34 % FEV1) exposure days, and progressively 

diminished toward baseline levels (!6 % FEV1) by the fifth exposure day. Similar to healthy 

subjects, asthmatics lose their tolerance to O3 after 4 to 7 days without O3 exposure. 

Some, but not all, studies have reported that asthmatics have a somewhat exaggerated 

inflammatory response to acute O3 exposure relative to healthy control subjects (e.g., McBride 

et al., 1994; Basha et al., 1994; Peden et al., 1995, 1997; Peden, 2001a; Scannell et al., 1996; 

Hiltermann et al., 1997, 1999; Michelson et al., 1999; Vagaggini et al., 1999; Newson et al., 

2000; Holz et al., 2002) see also Section 6.9 and Tables AX6-3 and AX6-12). For example, at 

18-h post-O3 exposure (0.2 ppm, 4-h with IE, V0 E = 25 L/min/m2

 BSA) and corrected for FA 

responses, Scannell et al. (1996) found significantly increased neutrophils in 18 asthmatics 

(12%) compared to 20 healthy subjects (4.5%). This inflammatory response difference was 

observed despite no group differences in spirometric responses to O3. Inflammatory responses 

do not appear to be correlated with lung function responses in either asthmatic or healthy 

subjects (Balmes et al., 1996, 1997; Holz et al., 1999). The lack of correlation between 

inflammatory and spirometric responses may be due to differences in the time kinetics of these 

different types of responses (Stenfors et al., 2002). In addition, airway responsiveness to 

inhaled allergens is increased by O3 exposure in subjects with allergic asthma for up to 24 h 

(see Section 6.8). 

6.3.3 Subjects with Allergic Rhinitis 

Allergic rhinitis is a condition defined by inflammation of the nasal membranes. Nayak 

(2003) recently reviewed the commonalities between asthma and allergic rhinitis. Clinically, 

greater than 60% of asthmatics have allergic rhinitis and slightly less than 40% of allergic 

rhinitics have asthma. Leukotrienes and histamine are well-recognized mediators of responses 

(viz., inflammation, hyperresponsiveness, and bronchoconstriction) in both asthma and allergic 

6-18
rhinitis. Although, rhinitis and asthma are distinguished as affecting the upper and lower 

airways, respectively, it has been suggested that these diseases are manifestations of a similar 

disease process. 

Given the prevalence of concomitant asthma and rhinitis and their common response 

mediators, it should be expected that allergic rhinitics might respond more similarly to 

asthmatics than healthy individuals. Regarding spirometric responses, Jörres et al. (1996) 

provide the only data demonstrating a trend in support of this supposition. 

Studies demonstrating the interaction between air pollutants and allergic processes in the 

human nasal airways and rhinoconjunctival tissue have been reviewed by Peden (2001b) and 

Riediker et al. (2001), respectively. Ozone exposure of subjects with allergic rhinitis has been 

shown to induce nasal inflammation and increase airway responsiveness to nonspecific 

bronchoconstrictors. 

Peden et al. (1995), who studied allergic asthmatics exposed to O3, found that O3 causes an 

increased response to nasal allergen challenge in addition to nasal inflammatory responses. 

Their data suggested that allergic subjects have an increased immediate response to allergen 

after O3 exposure. In a follow-up study, Michelson et al. (1999) reported that 0.4 ppm O3 did not 

promote early-phase-response mediator release or enhance the response to allergen challenge in 

the nasal airways of mild, asymptomatic dust mite-sensitive asthmatic subjects. Ozone did, 

however, promote an inflammatory cell influx, which helps induce a more significant late-phase 

response in this population. 

Jörres et al. (1996) found that O3 causes an increased response to bronchial allergen 

challenge in subjects with allergic rhinitis. This study also measured responses in healthy 

subjects and mildly allergic asthmatics (see Sections AX6.3.2 and AX6.8). All subjects were 

exposed to 0.25 ppm O3 for 3 h with IE ( V0 E = 30 L/min). Statistically significant O3-induced 

decrements in FEV1 occurred in rhinitics (14.1%), asthmatics (12.5%), and the healthy controls 

(10.2%), but these responses did not differ statistically between groups. Methacholine 

responsiveness was significantly increased in asthmatics, but not in subjects with allergic 

rhinitis. Airway responsiveness to an individual’s historical allergen (grass and birch pollen, 

house dust mite, or animal dander) was significantly increased after O3 exposure when compared 

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to FA exposure. The authors concluded that subjects with allergic rhinitis, but without asthma, 

could be at risk if a high O3 exposure is followed by a high dose of allergen. 

Holz et al. (2002) extended the results of Jörres et al. (1996) by demonstrating that 

repeated daily exposure to lower concentrations of O3 (0.125 ppm for 4 days) causes an 

increased response to bronchial allergen challenge in subjects with preexisting allergic airway 

disease, with or without asthma. These investigators observed no major difference in the pattern 

of bronchial allergen response between asthmatics or rhinitics, except that the dose of allergen 

required to elicit a similar response ($20% decrease in FEV1) was significantly greater for 

rhinitic than asthmatic subjects (16 vs. 0.6 mg/mL). Early phase responses were more consistent 

in subjects with rhinitis, and late-phase responses were more pronounced in subjects with 

asthma. There also was a tendency towards a greater effect of O3 in subjects with greater 

baseline response to specific allergens (chosen on the basis of skin prick test and history, viz., 

grass, rye, birch, or alder pollen, house dust mite, or animal dander). These data suggest that the 

presence of allergic bronchial sensitization, but not a history of asthma, may be a key 

determinant of increased airway allergen responsiveness following exposure to O3 (for a more 

complete discussion of airway responsiveness, see Section AX6.8). 

6.3.4 Subjects with Cardiovascular Disease 

Possibly due to the age of subjects studied, O3 exposure does not appear to result in 

significant pulmonary function impairment or evidence of cardiovascular strain in patients with 

cardiovascular disease relative to healthy controls. Gong et al. (1998) exposed 10 hypertensive 

and 6 healthy adult males, 41 to 78 years of age, to 0.3 ppm O3 for 3 h with IE at 30 L/min. For 

all subjects combined (no significant group differences), there was an O3-induced decrement of 

7% in FEV1 and an 70% increase in the alveolar-arterial oxygen tension gradient. The overall 

results did not indicate any major acute cardiovascular effects of O3 in either the hypertensive or 

normal subjects. Gong et al. (1998) suggested that by impairing alveolar-arterial oxygen 

transfer, the O3 exposure could potentially lead to adverse cardiac events by decreasing oxygen 

supply to the myocardium. However, the subjects in their study apparently had sufficient 

functional reserve so as to not experience significant ECG changes or myocardial ischemia 

and/or injury (see Section 6.10 for additional discussion). 

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6.4 INTERSUBJECT VARIABILITY AND REPRODUCIBILITY 

OF RESPONSE 

Analysis of factors that contribute to intersubject variability is important for the 

understanding of individual responses, mechanisms of response, and health risks associated with 

acute O3 exposures. Large intersubject variability in response to O3 has been reported by 

numerous investigators (Adams et al., 1981; Aris et al., 1995; Folinsbee et al., 1978; Kulle et al., 

1985; McDonnell et al., 1983). The magnitude of individual variability in FEV1 response in 2 h 

IE exposures increases at higher O3 concentrations (Kulle et al., 1985; McDonnell et al., 1983). 

McDonnell (1996) examined the FEV1 response data from three 6.6-h exposure studies 

conducted at the EPA Health Effects Research Laboratory and showed that the FEV1 responses 

in FA were small, with most tightly grouped around zero. With increasing O3 concentrations 

between 0.08 and 0.12 ppm, the mean response became asymmetrical, with a few individuals 

experiencing quite large decrements in FEV1 (intersubject variability observed in O3 dosimetry 

studies is discussed in Chapter 4, Section 4.2). 

As an example of the variation in spirometric responses to O3 exposure, Hazucha et al. 

(2003) analyzed the distribution of O3 responsiveness in 240 subjects (18 to 60 years of age) 

exposed to 0.42 ppm O3 (on 3 occasions) for 1.5 h with IE at V0 E = 20 L/min/m2

 BSA. Across 

all ages, 18% of subjects were weak responders (#5% FEV1 decrement), 39% were moderate 

responders, and 43% were strong responders ($15% FEV1 decrement). Younger subjects 

(#35 years of age) were predominately strong responders, whereas older subjects (>35 years of 

age) were mainly weak responders. The influence of age on intersubject variability was also 

noted by Passannante et al. (1998), who found that subjects under 35 years of age were more like 

to be strong responders than older individuals. 

For repeated exposures, Hazucha et al. (2003) reported that the reproducibility of FEV1 

responses was related to the length of time between exposures. A Spearman correlation 

coefficient of 0.54 was found between responses for exposures separated by 105 days (median), 

whereas a correlation coefficient of 0.85 was found between responses for exposures separated 

by only 7 days (median). The more reproducible the subject’s response, the more precisely it 

indicates his/her intrinsic responsiveness. In 2 h IE O3 exposures, McDonnell et al. (1985) found 

a relatively poor FEV1 reproducibility (R = 0.58) at the lowest concentration, 0.12 ppm, due in 

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part to a lack of specific O3 response or a uniformly small response in the majority of subjects. 

It was concluded that, for 2 h IE O3 exposures equal to or greater than 0.18 ppm, the intersubject 

differences in magnitude of change in FVC and FEV1 are quite reproducible over time (21 to 

385 days; mean = 33 days) and are due primarily to differences in intrinsic responsiveness of 

individual subjects to O3 exposure. 

Intersubject variability, mechanisms of response, and health risks associated with acute O3 

exposures are complicated by weak associations between various O3-induced responses. In a 

retrospective analysis of data for 485 male subjects (aged 18 to 36 yrs) exposed to one of six O3 

concentrations at one of three activity levels for 2 h, McDonnell et al. (1999) found significant, 

but low, Spearman rank order correlations between FEV1 response and symptoms of cough 

(R = 0.39), shortness of breath (R = 0.41), and pain on deep inspiration (R = 0.30). These 

authors concluded that these different responses are mechanistically related to some degree, but 

indicated that there does not appear to be a single factor which is responsible for the observed 

individual differences in O3 responsiveness across the spectrum of symptom and lung function 

responses. 

The effect of large intersubject variability on the ability to predict individual 

responsiveness to O3 was demonstrated by McDonnell et al. (1993). These investigators 

analyzed data for 290 male subjects (18 to 32 yrs of age) who underwent repeat 2 h IE exposures 

to one or more O3 concentrations ranging from 0.12 to 0.40 ppm. They attempted to identify 

personal characteristics (i.e., age, height, baseline pulmonary function, presence of allergies, and 

past smoking history) that might predict individual differences in FEV1 response. Only age 

contributed significantly to intersubject responsiveness (younger subjects were more 

responsive), accounting for just 4% of the observed variance. Interestingly, O3 concentration 

accounted for only 31% of the variance, strongly suggesting the importance of as yet undefined 

individual characteristics that determine FEV1 responsiveness to O3. The authors concluded that 

much individual variability in FEV1 response to O3 remains unexplained. 

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6.5 FACTORS MODIFYING RESPONSIVENESS TO OZONE 

6.5.1 Influence of Age 

Children, adolescents, and young adults (<18 yrs of age) appear, on average, to have nearly 

equivalent spirometric responses to O3, but have greater responses than middle-aged and older 

adults when exposed to comparable O3 doses (U.S. Environmental Protection Agency, 1996). 

Symptomatic responses to O3 exposure, however, appear to increase with age until early 

adulthood and then gradually decrease with increasing age (U.S. Environmental Protection 

Agency, 1996). In contrast to young adults, the diminished symptomatic responses in children 

and the elderly may put the latter groups at increased risk for continued O3 exposure. Although 

no new laboratory studies investigating O3 responses in children have been published since the 

1996 O3 AQCD, the epidemiological studies published during the last decade (see Section 

7.2.3.1 for details) are generally in agreement with the earlier laboratory studies. 

The ensuing discussion in this section provides information on average FEV1 responses to 

O3 exposure as a function of age in healthy adults ranging from 18 to 50 years of age. Beyond 

approximately 18 years of age, spirometric and symptom responses to O3 exposure begin to 

decline with increasing age. In healthy individuals, the rate of decline in O3 responsiveness 

appears to be greater in younger (18 to 35 yrs) versus middle aged (35 to 55 yrs) individuals 

(Passannante et al., 1998; Hazucha et al., 2003). Beyond this age (>55 yrs), acute O3 exposure 

elicits minimal spirometric changes. An average FEV1 decrement of ~3% has been reported by 

Gong et al. (1997a) for this older population under a “worst case” exposure scenario (0.24 ppm 

O3 with 4 h IE). Although Gong et al. (1997a) and others have examined responses to O3 

exposure in subjects of various ages, the exposure conditions differ between most studies in such 

a manner so that age effects remain uncertain. 

Three recent studies, which analyzed large data sets ($240 subjects) of similarly exposed 

subjects, show clearly discernable changes in FEV1 responses to O3 as a function of age. In one, 

Seal et al. (1996) analyzed O3-induced spirometric responses in 371 young nonsmokers (aged 

18 to 35 yrs) exposed for 2.3 h during IE at a V0 E of 25 L/min/m2

 BSA. On average, for the same 

O3 concentration (C), the response of 25-, 30-, and 35-yr old individuals are predicted to be 83, 

65, and 48%, respectively, of the response in a 20-yr old. For example, a 5.4% decrement in 

FEV1 is predicted for a 20-yr old exposed to 0.12 ppm O3 for 2.3 h IE ( V0 E = 25 L/min/m2 BSA), 

whereas, a similarly exposed 35-yr old is predicted to have only a 2.6% decrement. 

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McDonnell et al. (1997) examined FEV1 responses in 485 healthy white males (aged 18 to 

36 yrs) exposed once for 2 h to an O3 concentration of 0.0, 0.12, 0.18, 0.24, 0.30, or 0.40 ppm at 

rest or one of two levels of IE ( V0 E of 25 and 35 L/min/m2 BSA). For the same exposure 

conditions (C, V0 E , and duration), the average responses of 25-, 30-, and 35-yr old individuals are 

predicted to be 69, 48, and 33%, respectively, of the response in 20-yr olds. Hazucha et al. 

(2003) analyzed the distribution of O3 responsiveness in 240 subjects (18 to 60 years of age) 

exposed to 0.42 ppm O3 for 1.5 h with IE at V0 E = 20 L/min/m2 BSA. In males, the FEV1 

responses of 25-, 35-, and 50-yr olds are predicted to be 94, 83, and 50%, respectively, of the 

average response in 20-yr old males. In females, the FEV1 responses of 25-, 35-, and 50-yr olds 

are predicted to be 82, 46, and 18%, respectively, of the average response in 20-yr old females. 

For subjects aged 18 to 36 yrs, McDonnell et al. (1999) recently reported that symptom 

responses from O3 exposure also decrease with increasing age. Whether the same age-dependent 

pattern of O3 sensitivity decline also holds for airway reactivity or inflammatory endpoints has 

not been determined. 

6.5.2 Gender and Hormonal Influences 

Several studies have suggested that physiological differences between the genders may 

predispose females to a greater susceptibility to O3. Housley et al. (1996) reported that females 

have lower concentration of uric acid (the most prevalent antioxidant) in nasal lavage fluid than 

males. Since the source of uric acid (the plasma) is also known to be lower in females, the 

authors suggested that this could be a contributing factor leading to gender differences in 

response to oxidants. Plausibly, reduced absorption of O3 in the upper airways may promote its 

deeper penetration into the lower respiratory system regions. Dosimetric measurements have 

shown that the absorption distribution of O3 is independent of gender when absorption is 

normalized to anatomical dead space (Bush et al., 1996). More recently, Ultman et al. (2004) 

reported that the whole lung uptake fraction of O3 was significantly greater in males (91.4%) 

than females (87.1%). But, this increase in O3 uptake in the males was consistent with their 

larger tidal volume and slower breathing frequency relative to the females. Furthermore, O3 

uptake was not correlated with spirometric responses. Thus, a differential removal of O3 by uric 

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acid seems to have minimal effect. In general, the spirometric responses of healthy young 

females to O3 exposure appears comparable to the responses of young males (Hazucha et al., 

2003). Although, during the follicular phase of the menstrual cycle, lung function response to O3 

is enhanced (Fox et al., 1993). 

6.5.3 Racial, Ethnic, and Socioeconomic Status Factors 

In the only laboratory study designed to compare spirometric responses of whites and 

blacks exposed to a range of O3 concentrations (0 to 0.4 ppm), Seal et al. (1993) reported 

inconsistent and statistically insignificant FEV1 differences between white and black males and 

females within various exposure levels. Perhaps with larger cohorts, the tendency for greater 

responses of black than white males may become significant. Thus, based on this study it is still 

unclear if race is a modifier of O3 sensitivity, although the findings of epidemiologic studies 

reported in the previous, 1996 O3 AQCD “can be considered suggestive of an ethnic difference” 

(U.S. Environmental Protection Agency, 1996). 

Gwynn and Thurston (2001) pointed out, it appears that it is more the socioeconomic status 

(SES) and overall quality of healthcare that drives PM10- and O3-related hospital admissions than 

an innate or acquired sensitivity to pollutants. This assertion appears to be supported somewhat 

by the study of Seal et al. (1996), who employed a family history questionnaire to examine the 

influence of SES on the O3 responsiveness. Seal et al. found that, of the three SES categories, 

individuals in the middle SES category showed greater concentration-dependent decline in % 

predicted FEV1 (4 to 5% at 0.4 ppm O3) than low and high SES groups. The authors did not 

have an “immediately clear” explanation for this finding. With such a paucity of data, it is not 

possible to discern the influence of racial or economic-related factors on O3 sensitivity. 

6.5.4 Influence of Physical Activity 

Any physical activity will increase minute ventilation and therefore the dose of inhaled O3. 

Consequently, the intensity of physiological response following an acute exposure will be 

strongly associated with minute ventilation (see Figures 6-3 and AX6-2). 

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Figure 6-3. Predicted O3-induced decrements in FEV1 as a function of exposure duration 

and level of IE (line labels are V0 E levels) in healthy young adults (20 yrs of 

age) exposed to 0.3 ppm O3. The illustrated activity levels range from rest 

(VE 0 = 10 L/min) to moderate exercise ( V0 E = 40 L/min). Predictions are for 

Model 1 coefficients in Table 3 of McDonnell et al. (1997). 

Source: Based on McDonnell et al. (1997). 

6.5.5 Environmental Factors 

Since the 1996 O3 AQCD (U.S. Environmental Protection Agency, 1996) few human 

laboratory studies have examined the potential influence of environmental factors (e.g., rural 

versus urban environment, passive cigarette smoke exposure, and bioactive agents such as 

endotoxin) on O3-induced pulmonary function changes in healthy individuals. 

New controlled human exposure studies have confirmed that smokers are less responsive 

to O3 than nonsmokers. Spirometric and plethysmographic pulmonary function decline, 

nonspecific airway hyperreactivity, and inflammatory response of smokers to O3 were all weaker 

than those reported for nonsmokers. Although all these responses are intrinsically related, the 

functional association between them, as in nonsmokers, has been weak. Similarly, the time 

course of development and recovery of these effects, as well their reproducibility, was not 

different from nonsmokers. Chronic airway inflammation with desensitization of bronchial 

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nerve endings and an increased production of mucus may plausibly explain the smaller responses 

to O3 observed in smokers (Frampton et al., 1997; Torres et al., 1997). 

The effect of environmental tobacco smoke (ETS) on O3 responses has received very little 

attention. In one study (Yu et al., 2002), preexposure of mice to sidestream cigarette smoke 

(ETS surrogate) elicited no immediate effects, but potentiated subsequent O3-induced 

inflammatory responses (see Chapter 5.4.2 for additional ETS details). Endotoxin is a 

biologically active component of both mainstream and sidestream tobacco smoke (Hasday et al., 

1999), which might contribute to the potentiation of O3 effects. 

The influence of ambient temperature on pulmonary effects induced by O3 exposure in 

humans has been studied infrequently under controlled laboratory conditions. Several 

experimental human studies have reported additive effects of heat and O3 exposure (see U.S. 

Environmental Protection Agency, 1986, 1996). Foster et al. (2000) exposed 9 young healthy 

subjects for 130 min (IE 10 min at 36 to 39 l/min) to filtered air and to ramp profile O3 at 22 °C 

and 30 °C, 45-55% RH. The O3 exposure started at 0.12 ppm, reached a peak of 0.24 ppm 

midway through, and subsequently declined to 0.12 ppm at the end of exposure. At the end of 

exposure, FEV1 had decreased significantly (p < 0.5) by ~8% at 22 °C and ~6.5% at 30 °C. 

One day (19 h) later, the decline of 2.3% from baseline was still significant (p < 0.05) at both 

temperatures. FVC decrements were smaller and significant only for the 22 °C condition 

immediately postexposure. There was a decline in specific airway conductance (sGaw; p < 0.05) 

at 30 °C, but not at 22 °C. Nonspecific bronchial responsiveness to methacloline assessed 

as PC50 sGaw was significantly (p < 0.05) higher one day following O3 exposure at both 

temperatures, but more so at 30 °C. Thus, these findings suggest that elevated temperature may 

partially attenuate spirometric responses but enhance airway reactivity. 

6.5.6 Oxidant-Antioxidant Balance 

The first line of defense against oxidative stress involves antioxidants present in epithelial 

lining fluid (ELF), which scavenge free radicals and limit lipid peroxidation. Exposure to O3 

depletes the antioxidant level in nasal ELF, probably due to scrubbing of O3 (Mudway et al., 

1999); however, the concentration and the activity of antioxidant enzymes either in ELF or 

plasma do not appear to be related to O3 responsiveness (Avissar et al., 2000; Blomberg et al., 

1999; Samet et al., 2001). Carefully controlled studies of dietary antioxidant supplementation 

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have shown some protective effects of "-tocopherol and ascorbate on spirometric lung function 

from O3 but not on the intensity of subjective symptoms and inflammatory response, including 

cell recruitment, activation and release of mediators (Samet et al., 2001; Trenga et al., 2001). 

Dietary antioxidants have also afforded partial protection to asthmatics by attenuating 

postexposure bronchial hyperresponsiveness (Trenga et al., 2001). Field studies performed in 

Mexico City (described in Section 7.2.3.1) and animal studies (described in Section 5.2.1.3) have 

also demonstrated protective effects of ELF antioxidants during O3 exposures. 

6.5.7 Genetic Factors 

Several recent studies (Bergamaschi et al., 2001; Corradi et al., 2002; David et al., 2003; 

Romieu et al., 2004; Yang et al., 2005) have reported that genetic polymorphism of antioxidant 

enzymes and inflammatory genes may modulate pulmonary function and inflammatory response 

to O3 challenge. It has been suggested that healthy carriers of NAD(P)H:quinone oxidoreductase 

wild type (NQO1wt) in combination with glutathione S-transferase μ-1 genetic deficiency 

(GSTM1null) are more responsive to O3 (Bergamaschi et al., 2001). The authors reported that 

subjects having NQO1wt and GSTM1null genotypes had increased O3 responsiveness (FEV1 

decrements and epithelial permeability), whereas subjects with other combinations of these 

genotypes were less affected. A subsequent study from the same laboratory reported a positive 

association between O3 responsiveness, as characterized by the level of oxidative stress and 

inflammatory mediators (8-isoprostane, LTB4 and TBARS) in exhaled breath condensate and the 

NQO1wt and GSTM1null genotypes (Corradi et al., 2002). However, none of the spirometric 

lung function endpoints were affected by O3 exposure. It is of interest to note that human nasal 

mucosa biopsies of GSTM1null subjects showed higher antioxidant enzymes activity than 

biopsies of GSTM1 positive individuals when exposed to ozone (Otto-Knapp et al., 2003). 

Asthmatic children with a genetic deficiency of GSTM1 were reported to be more 

responsive to ambient O3 exposure than asthmatic children without this deficiency, as assessed 

by decrements in FEF25-75, in this field study (Romieu et al., 2004). Antioxidant supplementation 

(vit. C and E) attenuated post-O3 exposure lung function response in these children. More 

specific genotyping has shown that O3 responsiveness of asthmatic children may be related to the 

presence of variant Ser allele for NQO1. The presence of at least one NQO1 Ser allele in 

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combination with GSTM1 null genotype lowered the risk of asthma in O3-exposed asthmatic 

children relative to Pro/Pro genotype (David et al., 2003). 

The influence of functional polymorphism in TNF-", lymphotoxin- " (LTA), TLR4, SOD2 

and GPX1 genes on ozone-induced lung function changes in healthy individuals, mild asthmatics 

and subjects with rhinitis was varied. Of the inflammatory genes studied, only TNF-" has 

appeared to show some promise as one of the genetic factors of susceptibility. However, as the 

authors stated, “the functional significance of individual TNF-" polymorphisms remains 

controversial” (Yang et al., 2005). 

These recent studies have shown that an individual’s innate susceptibility to O3 may be 

linked to the genetic background of an individual. Although a number of potential O3 

susceptibility genes have been identified, additional better designed and controlled studies are 

needed to ascertain the link between susceptibility and specific polymorphism. 

6.6 REPEATED O3 EXPOSURE EFFECTS 

Based on studies reviewed here and in the previous O3 criteria documents (U.S. 

Environmental Protection Agency, 1986, 1996), several conclusions can be drawn about 

repeated 1- to 2-h O3 exposures. Repeated exposures to O3 can cause an enhanced (i.e., greater) 

pulmonary function response on the second day of exposure (see Tables AX6-8 and AX6-9 for 

added detail). This enhancement appears to be dependent on the interval between the exposures 

(24 h is associated with the greatest increase) and is absent with intervals greater than 3 days 

(Bedi et al., 1985; Folinsbee and Horvath, 1986; Schonfeld et al., 1989). An enhanced response 

also appears to depend to some extent on the magnitude of the initial response (Horvath et al., 

1981). Small responses to the first O3 exposure are less likely to result in an enhanced response 

on the second day of O3 exposure (Folinsbee et al., 1994). With continued daily exposures (i.e., 

beyond the second day) there is an attenuation of pulmonary function responses, typically after 

3 to 5 days of repeated O3 exposure. This attenuation of responses is lost in 1 week (Kulle et al., 

1982; Linn et al., 1982b) or perhaps 2 weeks (Horvath et al., 1981) without O3 exposure. 

In temporal conjunction with pulmonary function changes, symptoms induced by O3 (e.g., 

cough, pain on deep inspiration, and chest discomfort), are increased on the second exposure day 

and attenuated with repeated O3 exposure thereafter (Folinsbee et al., 1980, 1998; Foxcroft and 

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Adams, 1986; Linn et al., 1982b). O3-induced changes in airway responsiveness persist longer 

and attenuate more slowly than pulmonary function and symptoms responses (Dimeo et al., 

1981; Kulle et al., 1982), although this has been studied only on a limited basis (Folinsbee et al., 

1994). In longer-duration (4 h to 6.6 h), lower-concentration studies that do not cause an 

enhanced second-day response, the attenuation of response to O3 appears to proceed more 

rapidly (Folinsbee et al., 1994) (Effects of repeated exposures on inflammatory responses are 

discussed in Section 6.9.4). 

6.7 EFFECTS ON EXERCISE PERFORMANCE 

The effects of acute O3 inhalation on endurance exercise performance have been examined 

in numerous controlled laboratory studies. These studies were discussed in the 1996 O3 AQCD 

(U.S. Environmental Protection Agency, 1996) and can be divided into two categories: (1) those 

that examined the effects of acute O3 inhalation on maximal oxygen uptake ( VO0 2max) and 

(2) those that examined the effects of acute O3 inhalation on the ability to complete strenuous 

continuous exercise protocols of up to 1 h in duration. 

In brief, endurance exercise performance and VO0 2max may be limited by acute exposure 

to O3 (Adams and Schelegle, 1983; Schelegle and Adams, 1986; Gong et al., 1986; Foxcroft and 

Adams, 1986; Folinsbee et al., 1977; Linder et al., 1988). Gong et al. (1986) and Schelegle and 

Adams (1986) found that significant reductions in maximal endurance exercise performance may 

occur in well-conditioned athletes while they perform CE ( V0 E > 80 L/min) for 1 h at O3 

concentrations $0.18 ppm. Reports from studies of exposure to O3 during high-intensity 

exercise indicate that breathing discomfort associated with maximal ventilation may be an 

important factor in limiting exercise performance in some, but not all, subjects. 

6.8 EFFECTS ON AIRWAY RESPONSIVENESS 

Airway hyperresponsiveness refers to a condition in which the propensity for the airways 

to bronchoconstrict due to a variety of stimuli becomes augmented. Airway responsiveness is 

typically quantified by measuring the decrement in pulmonary function (i.e., spirometry or 

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plethysmography) following inhalation of small amounts of an aerosolized specific (antigen, 

allergen) or nonspecific (methacholine, histamine) bronchoconstrictor agent or some other 

measured stimulus (e.g., exercise, cold air). 

Ozone exposure causes an increase in nonspecific airway responsiveness as indicated by a 

reduction in the concentration of methacholine or histamine required to produce a given 

reduction in FEV1 or increase in SRaw. Increased airway responsiveness is an important 

consequence of exposure to O3, because its presence means that the airways are predisposed to 

narrowing upon inhalation of a variety of stimuli (e.g., specific allergens, SO2, cold air). 

Ozone exposure of asthmatic subjects, who characteristically have increased airway 

responsiveness at baseline, can cause further increases in responsiveness (Kreit et al., 1989). 

Similar relative changes in airway responsiveness are seen in asthmatics exposed to O3 despite 

their markedly different baseline airway responsiveness. Several studies (Jörres et al., 1996; 

Kehrl et al., 1999; Molfino et al., 1991) have been published suggesting an increase in specific 

(i.e., allergen-induced) airway reactivity in response to O3 exposure. An important aspect of 

increased airway responsiveness after O3 exposure is that this represents a plausible link between 

ambient O3 exposure and increased hospital admissions for asthma. 

Changes in airway responsiveness after O3 exposure appear to be resolved more slowly 

than changes in FEV1 or respiratory symptoms (Folinsbee and Hazucha, 2000). Furthermore, in 

studies of repeated exposure to O3, changes in airway responsiveness tend to be somewhat less 

susceptible to attenuation with consecutive exposures than changes in FEV1 (Dimeo et al., 1981; 

Folinsbee et al., 1994; Gong et al., 1997b; Kulle et al., 1982). Ozone-induced increases in 

airway responsiveness do not appear to be strongly associated with decrements in lung function 

or increases in symptoms. 

Mechanisms underlying O3-induced increases in airway responsiveness are only partially 

understood, but such increases appear to be associated with a number of cellular and biochemical 

changes in airway tissue. Although inflammation could play a role in the increase in airway 

responsiveness, cyclooxygenase inhibitors have not been effective at blocking the O3-induced 

influx of PMNs into bronchoalveolar lavage (BAL) fluid (Hazucha et al., 1996; Ying et al., 

1990). Therefore, O3-induced airway responsiveness may not be due to the presence of PMNs 

in the airway or to the release of arachidonic acid metabolites. Rather, it seems likely that the 

mechanism for this response is multifactorial, possibly involving the presence of cytokines, 

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prostanoids, or neuropeptides; activation of macrophages, eosinophils, or mast cells; and 

epithelial damage that increases direct access of mediators to the smooth muscle or receptors in 

the airways that are responsible for reflex bronchoconstriction. 

6.9 EFFECTS ON INFLAMMATION AND HOST DEFENSE 

6.9.1 Introduction 

Short-term exposure of humans to O3 can cause acute inflammation, and long-term 

exposure of laboratory animals results in a chronic inflammatory state (see Chapter 5). The 

relationship between repetitive bouts of acute inflammation in humans caused by O3 and the 

development of chronic respiratory disease is unknown. 

The presence of neutrophils (PMNs) in the lung has long been accepted as a hallmark of 

inflammation and is an important indicator that O3 causes inflammation in the lungs. It is 

apparent, however, that inflammation within airway tissues may persist beyond the point that 

inflammatory cells are found in BAL fluid (BALF). Soluble mediators of inflammation such as 

the cytokines (IL-6, IL-8) and arachidonic acid metabolites (e.g., PGE2, PGF2", thromboxane, 

and leukotrienes [LTs] such as LTB4) have been measured in the BAL fluid of humans exposed 

to O3. In addition to their role in inflammation, many of these compounds have 

bronchoconstrictive properties and may be involved in increased airway responsiveness 

following O3 exposure. 

Some recent evidence suggests that changes in small airways function may provide a 

sensitive indicator of O3 exposure and effect, despite the fact that inherent variability in their 

measurement by standard spirometric approaches makes their assessment difficult (Frank et al., 

2001). Observations of increased functional responsiveness of these areas relative to the more 

central airways, and of persistent effects following repeated exposure, may indicate that further 

investigation of inflammatory processes in these regions is warranted. 

6.9.2 Inflammatory Responses in the Upper Respiratory Tract 

The nasal passages constitute the primary portal for inspired air at rest and, therefore, 

the first region of the respiratory tract to come in contact with airborne pollutants. Nikasinovic 

et al. (2003) recently reviewed the literature of laboratory-based nasal inflammatory studies 

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published since 1985. Nasal lavage (NL) has provided a useful tool for assessing O3-induced 

inflammation in the nasopharynx. Increased levels of PMNs in the NL fluid of humans exposed 

to 0.5 ppm O3 at rest for 4 h has been reported (Graham et al., 1988; Bascom et al., 1990). 

Graham and Koren (1990) compared inflammatory mediators present in both the NL and 

BAL fluids of humans exposed to 0.4 ppm O3 for 2 h. Similar increases in PMN were observed 

in NL and BAL, suggesting a qualitative correlation between inflammatory changes in the lower 

airways (BAL) and the upper respiratory tract (NL). Torres et al. (1997) compared NL and BAL 

in smokers and nonsmokers exposed to 0.22 ppm O3 for 4 h. In contrast to Graham and Koren 

(1990), they did not find a relationship between numbers or percentages of PMNs in the nose 

and the lung, perhaps in part due to the variability observed in their NL recoveries. Albumin, 

a marker of epithelial cell permeability, was increased 18 h later, but not immediately after 

exposure, as seen by Bascom et al. (1990). 

McBride et al. (1994) reported that asthmatic subjects were more sensitive than 

nonasthmatics to upper airway inflammation at an O3 concentration (0.24 ppm for 1.5 h with 

light IE) that did not affect pulmonary function. In the asthmatics, there was a significant 

increase in the number of PMNs in NL fluid both immediately and 24 h after exposure. Peden 

et al. (1995) also found that exposure to 0.4 ppm O3 had a direct nasal inflammatory effect and a 

priming effect on response to nasal allergen challenge. A subsequent study in dust 

mite-sensitive asthmatic subjects indicated that O3 at this concentration enhanced eosinophil 

influx in response to allergen but did not promote the early-phase nasal response to allergen 

(Michelson et al., 1999). Similar to observations made in the lower airways, the presence of O3 

molecular “targets” in nasal lining fluid is likely to provide some level of local protection against 

exposure. In a study of healthy subjects exposed to 0.2 ppm O3 for 2 h, Mudway and colleagues 

(1999) observed a significant depletion of uric acid in NL fluid at 1.5 h following exposure. 

6.9.3 Inflammatory Response in the Lower Respiratory Tract 

As reviewed in the 1996 O3 AQCD (U.S. Environmental Protection Agency, 1996), acute 

exposure to O3 results in an inflammatory reaction, increased epithelial cell permeability, and 

may stimulate fibrogenic processes. Inflammatory markers are observed in BALF of healthy 

subjects by 1 h post-O3 exposure and may persist for at least 18 to 24 h. Not all inflammatory 

6-33
markers, however, follow the same time course. Studies published since the 1996 O3 AQCD 

support these earlier findings. 

Inflammatory effects have been assessed in vivo by lavage (proximal airway and 

bronchoalveolar), bronchial biopsy, and more recently, induced sputum. In vitro studies of 

human alveolar macrophages (AM) and airway epithelial cells exposed to O3 suggest that these 

cell types release mediators found in the BALF of O3-exposed humans (U.S. Environmental 

Protection Agency, 1996). Recent evidence suggests that the release of mediators from AMs 

may be modulated by the products of O3-induced oxidation of airway lining fluid components, 

such as human surfactant protein A (Wang et al., 2002). 

Spirometric responses to O3 are independent of inflammatory responses and markers of 

epithelial injury (Balmes et al., 1996; Blomberg et al., 1999; Hazucha et al., 1996; Torres et al., 

1997). Significant inflammatory responses to O3 exposures that did not elicit significant 

spirometric responses have been reported (Holz et al., 2005; McBride et al., 1994). 

A meta-analysis of 21 studies (Mudway and Kelly, 2004) showed that PMN influx in healthy 

subjects is associated with total O3 dose (product of O3 concentration, exposure duration, 

and V0 E ). 

The time course of the inflammatory response to O3 in humans has not been fully 

characterized. From a review of the literature by Mudway and Kelly (2000), Figure 6-4 

illustrates a plausible time course of acute O3 responses. As the figure shows, different markers 

exhibit peak responses at different times. Studies in which lavages were performed 1 h after O3 

exposure (1 h at 0.4 ppm or 4 h at 0.2 ppm) have demonstrated that the inflammatory responses 

are quickly initiated (Devlin et al., 1996; Schelegle et al., 1991; Torres et al., 1997). 

Inflammatory mediators and cytokines such as IL-8, IL-6, and PGE2 are greater at 1 h than at 

18 h post-O3 exposure (Devlin et al., 1996; Torres et al., 1997). However, IL-8 still remain 

elevated at 18 h post-O3 (4 h at 0.2 ppm O3 versus FA) in healthy subjects (Balmes et al., 1996). 

Schelegle et al. (1991) found increased PMNs in the “proximal airway” lavage at 1, 6, and 24 h 

after O3 exposure (4 h at 0.2 ppm O3), with a peak response at 6 h. Although, at 18 to 24 h after 

O3 exposure, PMNs remain elevated relative to 1 h postexposure (Schelegle et al., 1991; Torres 

et al., 1997). In addition to the influx of PMNs, lymphocyte numbers in BALF are elevated 

significantly at 6 h following exposure (2 h at 0.2 ppm O3) of healthy subjects (Blomberg et al., 

1997). Flow cytometry also indicated the increased presence of CD3+, CD4+ and CD8+ T cell 

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Figure 6-4. Time course of acute responses seen in humans exposed to O3. Responses are 

divided into three phases: initiation, during O3 exposure; progression, where 

responses develop post-O3 exposure; and resolution, during which responses 

return to baseline levels. *The IL-8 response is shown, with a progressive 

increase peaking at 18 h postexposure (18h-PE). The IL-8 literature is 

inconclusive and neutrophils have been demonstrated in the absence of an 

IL-8 increase. **Few studies have measured MPO; however, a trend for an 

increase at 6h-PE has been reported. 

Abbreviations: AA (ascorbic acid), FN (fibronectin), GSH (reduced 

glutathione), IL (interleukin), LDH (lactate dehydrogenase), LOP (lipid 

ozonation products), LTB4 (leukotriene B4), LTC4 (leukotriene C4), MPO 

(myeloperoxidase), PGE2 (prostaglandin E2), UA (uric acid). 

Source: Reprinted from Mudway and Kelly (2000) with permission from Elsevier. 

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subsets. This same laboratory later demonstrated that within 1.5 h following exposure of healthy 

subjects to the same O3 regimen, expression of human leukocyte antigen (HLA)-DR on lavaged 

macrophages underwent a significant, 2.5-fold increase (Blomberg et al., 1999). 

The inflammatory responses to O3 exposure have also been studied in asthmatic subjects 

(Basha et al., 1994; Scannell et al., 1996; Peden et al., 1997). In these studies, asthmatics 

showed significantly more neutrophils in the BALF (18 h postexposure) than similarly exposed 

healthy individuals. In one of these studies (Peden et al., 1997), which included only allergic 

asthmatics who tested positive for Dematophagoides farinae antigen, there was an eosinophilic 

inflammation (2-fold increase), as well as neutrophilic inflammation (3-fold increase). In a 

study of subjects with intermittent asthma exposed to 0.4 ppm O3 for 2 h, increases in eosinophil 

cationic protein, neutrophil elastase and IL-8 were found to be significantly increased 16 h 

post-exposure and comparable in induced sputum and BALF (Hiltermann et al, 1999). Scannell 

et al. (1996) also reported that IL-8 tends to be higher in the BALF of asthmatics compared to 

nonasthmatics following O3 exposure, suggesting a possible mediator for the significantly 

increased neutrophilic inflammation in those subjects. Bosson et al. (2003) found significantly 

greater epithelial expression of IL-5, IL-8, granulocyte-macrophage colony-stimulating factor 

(GM-CSF) and epithelial cell-derived neutrophil-activating peptide 78 (ENA-78) in asthmatics 

compared to healthy subjects following exposure to 0.2 ppm O3 for 2 h. In contrast, Stenfors 

et al. (2002) did not detect a difference in the O3-induced increases in neutrophil numbers 

between 15 mild asthmatic and 15 healthy subjects by bronchial wash at the 6 h postexposure 

time point. However, the asthmatics were on average 5 years older than the healthy subjects in 

this study, and it is not yet known how age affects inflammatory responses. It is also possible 

that the time course of neutrophil influx differs between healthy and asthmatic individuals. 

Vagaggini et al. (2002) investigated the effect of prior allergen challenge on responses in 

mild asthmatics exposed for 2 h to 0.27 ppm O3 or filtered air. At 6 h postexposure, eosinophil 

numbers in induced sputum were found to be significantly greater after O3 than after air 

exposures. Studies such as these suggest that the time course of eosinophil and neutrophil influx 

following O3 exposure can occur at levels detectable within the airway lumen by as early as 6 h. 

They also suggest that the previous or concurrent activation of proinflammatory pathways within 

the airway epithelium may enhance the inflammatory effects of O3. For example, in an in vitro 

6-36
study of epithelial cells from the upper and lower respiratory tract, cytokine production induced 

by rhinovirus infection was enhanced synergistically by concurrent exposure to O3 at 0.2 ppm for 

3 h (Spannhake et al, 2002). 

Although the release of mediators has been demonstrated to occur at exposure times and 

concentrations that are minimally cytotoxic to airway cells, potentially detrimental latent effects 

have been demonstrated in the absence of cytotoxicity. These include the generation of DNA 

single strand breaks (Kozumbo et al., 1996), the loss of cellular replicative activity (Gabrielson 

et al., 1994) in bronchial epithelial cells exposed in vitro, and the formation of protein and 

DNA adducts. A highly toxic aldehyde formed during O3-induced lipid peroxidation is 4­

hydroxynonenal (HNE). Healthy human subjects exposed to 0.4 ppm O3 for 1 h underwent BAL 

6 h later. Analysis of lavaged AMs by Western Blot indicated increased levels of a 32-kDa 

HNE-protein adduct, as well as 72-kDa heat shock protein and ferritin in O3-versus air-exposed 

subjects (Hamilton et al., 1998). In a recent study of healthy subjects exposed to 0.1 ppm O3 for 

2 h (Corradi et al., 2002), formation of 8-hydroxy-2'-deoxyguanosine (8-OHdG), a biomarker 

of reactive oxidant species (ROS)-DNA interaction, was measured in peripheral blood 

lymphocytes. At 18 h post exposure, 8-OHdG was significantly increased in cells compared to 

preexposure levels, presumably linked to concurrent increases in chemical markers of ROS. Of 

interest, the increase in 8-OHdG was only significant in a subgroup of subjects with the wild 

genotype for NQO1 and the null genotype for GSTM1, suggesting that polymorphisms in 

antioxidant enzymes may confer “susceptibility’ to O3 in some individuals. 

The generation of ROS following exposure to O3 has been shown to be associated with a 

wide range of responses. In a recent study, ROS production by alveolar macrophages lavaged 

from subjects exposed to 0.22 ppm for 4 h was assessed by flow cytometry (Voter et al., 2001). 

Levels were found to be significantly elevated 18 h postexposure and to be associated with 

several markers of increased permeability. An in vitro study of human tracheal epithelial cells 

exposed to O3 indicated that generation of ROS resulted in decreased synthesis of the 

bronchodilatory prostaglandin, PGE2, as a result of inactivation of prostaglandin endoperoxide 

G/H synthase 2 (Alpert et al., 1997). 

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6.9.4 Adaptation of Inflammatory Responses 

Physiologic and symptomatic responses in humans following repeated exposure to O3 were 

discussed in Section 6.6. Inflammatory responses upon repeated O3 exposures are discussed in 

this section. Animal studies suggest that while inflammation may be diminished with repeated 

exposure, underlying damage to lung epithelial cells continues (Tepper et al., 1989). Markers 

from BALF following both 2-h (Devlin et al., 1997) and 4-h (Christian et al., 1998; Jörres et al., 

2000) repeated O3 exposures (up to 5 days) indicate that there is ongoing cellular damage 

irrespective of the attenuation of some cellular inflammatory responses of the airways, 

pulmonary function, and symptom responses. 

Devlin et al. (1997) examined the inflammatory responces of humans repeatedly exposed 

to 0.4 ppm O3 for 5 consecutive days. Several indicators of inflammation (e.g., PMN influx, 

IL-6, PGE2, fibronectin) were attenuated after 5 days of exposure (i.e., values were not different 

from FA). Several markers (LDH, IL-8, total protein, epithelial cells) did not show attenuation, 

indicating that tissue damage probably continues to occur during repeated exposure. The 

recovery of the inflammatory response occurred for some markers after 10 days, but some 

responses were not normalized even after 20 days. The continued presence of cellular injury 

markers indicates a persistent effect that may not necessarily be recognized due to the 

attenuation of spirometric and symptom responses. 

Christian et al. (1998) randomly subjected heathy subjects to a single exposure and to 

4 consecutive days of exposure to 0.2 ppm O3 for 4 h. Both “bronchial” and “alveolar” fractions 

of the BAL showed decreased numbers of PMNs and fibronectin concentration at day 4 versus 

the single exposure, and a decrease in IL-6 levels in the alveolar fraction. Following a similar 

study design and exposure parameters, Jörres et al. (2000) found both functional and BALF 

cellular responses to O3 were abolished at 24 h postexposure following the fourth exposure day. 

However, levels of total protein, IL-6, IL-8, reduced glutathione and ortho-tyrosine were still 

increased significantly. In addition, visual scores for bronchitis, erythema and the numbers of 

neutrophils in the mucosal biopsies were increased. Their results indicate that, despite reduction 

of some markers of inflammation in BALF and measures of large airway function, inflammation 

within the airways persists following repeated exposure to O3. 

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Holz et al. (2002) made a comparison of early and late responses to allergen challenge 

following O3 in subjects with allergic rhinitis or allergic asthma. With some variation, both early 

and late FEV1 and cellular responses in the two subject groups were significantly enhanced by 

4 consecutive days of exposure to 0.125 ppm O3 for 3 h. 

In another study, Frank and colleagues (2001) exposed healthy subjects to FA and to O3 

(0.25 ppm, 2 h) on 4 consecutive days each, with pulmonary function measurements being made 

prior to and following each exposure. BAL was performed on day 5, 24 h following the last 

exposure. On day 5, PMN numbers remained significantly higher following O3 compared to FA. 

Of particular note in this study was the observation that small airway function, assessed by 

grouping values for isovolumetric FEF25-75, Vmax50 and Vmax75 into a single value, showed 

persistent reduction from day 2 through day 5. These data suggest that techniques monitoring 

the function in the small peripheral airway regions (the primary sites of O3 uptake in the lung), 

may provide important information regarding both acute and cumulative effects of O3 exposure. 

6.9.5 Effect of Anti-Inflammatory and Other Mitigating Agents 

Pretreatment of healthy subjects with non-steroidal anti-inflammatory drugs (ibuprofen, 

etc.) has been found to partially suppress development of airway inflammation and pulmonary 

function changes (U.S. Environmental Protection Agency, 1996). Although atropine blocked the 

increase in Raw in response to O3 exposure, it did not alter the spirometric or symptom 

responses (Beckett et al., 1985). Similarly, albuterol and salbutamol, which had no effect 

on O3-induced changes in spirometry, also had no effect of symptom responses (McKenzie et al., 

1987; Gong et al., 1988). The anti-inflammatory medications indomethacin and ibuprofen, 

which partially inhibit the spirometric responses to O3 exposure, also cause a reduction in 

respiratory symptoms (Schelegle et al., 1987; Hazucha et al., 1994). Indomethacin attenuates 

decrements in FEV1 and FVC in healthy subjects, but not asthmatics (Alexis et al., 2000). 

In contrast, inhalation of the corticosteroid budesonide does not prevent or even 

attenuate O3-induced responses in healthy subjects as assessed by measurements of lung 

function, bronchial reactivity and airway inflammation (Nightingale et al., 2000). In asthmatic 

subjects, budesonide decreases airway neutrophil influx following O3 exposure (Vagaggini 

et al., 2001). 

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Holz et al. (2005) studied inflammatory responses in healthy ozone-responders (>10% 

increase in sputum neutrophils from O3) pretreated with single doses (the highest shown to be 

safe and well tolerated) of inhaled fluticasone and oral prednisolone. The O3 exposure caused 

small changes in FEV1 (!3.6% ± 6.8%) that were not significantly different from baseline or 

between treatment groups (i.e., prescreening, placebo, fluticasone, and prednisolone). Relative 

to placebo, the inhaled or oral corticosteroids significantly reduced O3-induced neutrophil levels. 

These authors noted that their study design was intended to test the anti-inflammatory effects of 

the steroids and that such high-dose regimens should not be considered for potential long-term 

patient treatment. 

6.9.6 Changes in Host Defense Capability Following Ozone Exposures 

A number of studies clearly show that a single acute exposure (1 to 4 h) of humans to 

moderate concentrations of O3 (0.2 to 0.6 ppm) while exercising at moderate to heavy levels 

results in a number of cellular and biochemical changes in the lung, including an inflammatory 

response characterized by increased numbers of PMNs, increased permeability of the epithelial 

cells lining the respiratory tract, cell damage, and production of proinflammatory cytokines and 

prostaglandins. These responses can be detected as early as 1 h after exposure (Koren et al., 

1991; Schelegle et al., 1991) and can persist for at least 18 h (Aris et al., 1993; Koren et al., 

1989). The response profile of these mediators is not defined adequately, although it is clear that 

the time course of response varies for different mediators and cells (Devlin et al., 1997; 

Schelegle et al., 1991). Following a 2 h exposure to 0.2 ppm O3, Blomberg et al. (2003) 

observed increased Clara cell protein (biomarker of epithelial permeability) in blood at 2 h 

postexposure, which remained high at 6 h and had returned to baseline by 18 h. These changes 

also occur in humans exposed to 0.08 and 0.10 ppm O3 for 6 to 8 h (Devlin et al., 1991; Peden 

et al., 1997). Ozone also causes inflammatory changes in the nose, as indicated by increased 

levels of PMNs and albumin, a marker for increased epithelial cell permeability. Nasal lavage 

analyses, however, are not necessarily parallel to BALF analyses. 

There appears to be no strong correlation between any of the measured cellular and 

biochemical changes and changes in lung function measurements, suggesting that different 

mechanisms may be responsible for these processes (Balmes et al., 1996; Devlin et al., 1991). 

The idea of different mechanisms is supported by a study in which ibuprofen, a cyclooxygenase 

6-40
inhibitor, blunted the O3-induced decrements in lung function without altering the O3-induced 

increase in PMNs or epithelial cell permeability (Hazucha et al., 1996). In vitro studies suggest 

that epithelial cells are the primary target of O3 in the lung and that O3 induces them to produce 

many of the mediators found in the BALF of humans exposed to O3. Although O3 does not 

induce AMs to produce these compounds in large quantities, it does directly impair the ability of 

AMs to phagocytize and kill microorganisms. 

A number of studies have found that O3 exposures increases epithelial cell permeability 

through direct (technetium-99m labeled diethylene triamine pentaacetic acid, 99mTc-DTPA, 

clearance) and indirect (e.g., increased BALF albumin, protein) techniques. Kehrl et al. (1987) 

showed increased 99mTc-DTPA clearance in healthy young adults at 75 minutes postexposure 

to 0.4 ppm O3 for 2 h. More recently, Foster and Stetkiewicz (1996) have shown that 

increased 99mTc-DTPA clearance persists for at least 18-20 h post-O3 exposure (130 min to 

average O3 concentration of 0.24 ppm), and the effect is greater at the lung apices than at the 

base. Increased BALF protein, suggesting O3-induced changes in epithelial permeability, have 

also been reported at 1 h and 18 h postexposure (Balmes et al., 1996; Devlin et al., 1997). 

A recent meta-analysis of results from 21 publications (Mudway and Kelly, 2004), showed that 

increased BALF protein is associated with total ozone dose (product of O3 concentration, 

exposure duration, and V0 E ). Changes in permeability associated with acute inflammation may 

provide increased access of inhaled antigens, particles, and other substances to the smooth 

muscle, interstitial cells, and the blood. 

In addition to affecting epithelial permeability and AM-mediated clearance in the 

respiratory region of the lung, mucociliary clearance of the tracheobronchial airways is also 

affected by O3 exposure. Only two studies (Foster et al., 1987; Gerrity et al., 1993) have 

investigated the effect of O3 exposure on mucociliary particle clearance in humans. Foster et al. 

(1987) measured clearance during and after a 2 h exposure to 0.4 ppm O3. Gerrity et al. (1993) 

measured clearance at 2 h postexposure (0.4 ppm O3), by which time sRaw had returned to 

baseline and FVC was within 5% of baseline (versus an 11% decrement immediately 

postexposure). Foster et al. (1987) found a stimulatory effect of acute O3 exposure on 

mucociliary clearance. Gerrity et al. (1993), who observed no effect on clearance, suggested that 

transient clearance increases are coincident to pulmonary function responses. Investigators in 

6-41
both studies suggested that O3-induced increases in mucociliary clearance could be mediated by 

cholinergic receptors. 

6.10 EXTRAPULMONARY EFFECTS OF OZONE 

Ozone reacts rapidly on contact with respiratory system tissue and is not absorbed or 

transported to extrapulmonary sites to any significant degree as such. Human exposure studies 

discussed in the previous O3 AQCDs (U.S. Environmental Protection Agency, 1986, 1996) failed 

to demonstrate any consistent extrapulmonary effects. More recently, some human exposure 

studies have attempted to identify specific markers of exposure to O3 in blood. Foster et al. 

(1996) found a reduction in the serum levels of the free radical scavenger "-tocopherol after O3 

exposure. Liu et al. (1997, 1999) used a salicylate metabolite, 2,3, dehydroxybenzoic acid 

(DHBA), to indicate increased levels of hydroxyl radical which hydroxylates salicylate to 

DHBA. Increased DHBA levels after exposure to 0.12 and 0.40 ppm suggest that O3 increases 

production of hydroxyl radical. The levels of DHBA were correlated with spirometry changes. 

Relative to preexposure, Corradi et al. (2002) observed increased levels of 8-OHdG (a biomarker 

ROS-DNA interaction) in peripheral blood lymphocytes of healthy subjects at 18 h postexposure 

to 0.1 ppm O3 for 2 h. 

Gong et al. (1998) observed a statistically significant O3-induced increase the alveolar-toarterial PO2 gradient in both healthy (n = 6) and hypertensive (n = 10) adult males (aged 41 to 

78 yrs) exposed for 3 h with IE ( V0 E . 30 L/min) to 0.3 ppm O3. The mechanism for the 

decrease in arterial oxygen tension in the Gong et al. (1998) study could be due to an O3-induced 

ventilation-perfusion mismatch. Foster et al. (1993) has demonstrated that even in relatively 

young healthy adults (26.7 ± 7 yrs old), O3 exposure can cause ventilation to shift away from the 

well perfused basal lung. This effect of O3 on ventilation distribution (and, by association, the 

small airways) may persist beyond 24-h postexposure (Foster et al., 1997). Gong et al. (1998) 

suggested that by impairing alveolar-arterial oxygen transfer, the O3 exposure could potentially 

lead to adverse cardiac events by decreasing oxygen supply to the myocardium. The subjects in 

the Gong et al. (1998) study had sufficient functional reserve so as to not experience significant 

ECG changes or myocardial ischemia and/or injury. 

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Effects of O3 exposure on alveolar-arterial oxygen gradients may be more pronounced in 

patients with preexisting obstructive lung diseases. Relative to healthy elderly subjects, COPD 

patients have reduced gas exchange and low SaO2. Any inflammatory or edematous responses 

due to O3 delivered to the well-ventilated regions of the COPD lung could further inhibit gas 

exchange and reduce oxygen saturation. In addition, O3-induced vasoconstriction could also 

acutely induce pulmonary hypertension. Inducing pulmonary vasoconstriction and hypertension 

in these patients would perhaps worsen their condition, especially if their right ventricular 

function was already compromised. 

6.11 EFFECTS OF OZONE MIXED WITH OTHER POLLUTANTS 

Over the past 10 years only a handful of controlled-exposure human studies have examined 

the effects of pollutant mixtures containing O3. The studies summarized in this section 

complement the studies reviewed in the 1996 O3 AQCD (U.S. Environmental Protection 

Agency, 1996). (The complexities of O3 and co-pollutant exposures in animal studies are 

discussed in Chapter 5, Section 5.4.4). 

The results of a controlled study on children (Linn et al., 1997), designed to approximate 

exposure conditions of an epidemiologic study (Neas et al., 1995) by matching the population 

and exposure atmosphere (0.1 ppm O3, 0.1 ppm SO2 and 101 μg/m2 h2SO4), did not support the 

findings of this epidemiologic study. The study points out the difficulties in attempting to link 

the outcomes of epidemiologic and controlled studies. Another vulnerable population, 

asthmatics, demonstrated enhanced airway reactivity to house dust mite following exposures 

to O3, NO2, and the combination of the two gases. Spirometric response, however, was impaired 

only by O3, and O3 + NO2 at higher concentrations (Jenkins et al., 1999). Continuous exposure 

to SO2 and NO2 increases inhaled bolus O3 absorption, while continuous exposure to O3 

decreases O3 bolus absorption (Rigas et al., 1997). Inhalation of a mixture of PM2.5 and O3 by 

healthy subjects increased brachial artery tone and reactivity (Brook et al., 2002). Since no other 

cardiovascular endpoints were affected by the exposure, the pathophysiological importance of 

this observation remains uncertain. However, acute pulmonary hypertension due to O3-induced 

vasoconstriction could pose a risk to individuals with cardiovascular disease (see Section 6-10). 

6-43
All in all, the contention that air pollutant mixtures elicit stronger pathophysiologic effects 

than individual pollutants of the mix is only weakly supported by human studies of either healthy 

or at-risk individuals. 

6.12 CONTROLLED STUDIES OF AMBIENT AIR EXPOSURES 

A large amount of informative O3 exposure-effects data has been obtained in controlled 

exposure studies of humans or laboratory animals under a variety of different experimental 

conditions. However, laboratory simulation of the variable pollutant mixtures present in ambient 

air is not practical. Thus, the exposure effects of one or several artificially generated pollutants 

(i.e., a simple mixture) on pulmonary function and symptoms may not fully explain responses to 

ambient air exposures where complex pollutant mixtures exist. 

6.12.1 Mobile Laboratory Studies 

Quantitatively useful information on the effects of acute exposure to photochemical 

oxidants on pulmonary function responses and symptoms derived from field studies using a 

mobile laboratory were presented in prior criteria documents (U.S. Environmental Protection 

Agency, 1986, 1996). Relative to controlled exposure studies, mobile laboratory ambient air 

studies suffer an additional limitation in terms of a dependence on outdoor ambient conditions. 

Consistent with controlled exposure studies, mobile studies in California demonstrated that 

pulmonary effects from exposure to ambient air in Los Angeles are related to O3 concentration 

and level of exercise. Healthy subjects with a history of allergy also appeared to be more 

responsive to O3 than “nonallergic” subjects (Linn et al., 1980, 1983b), although a standardized 

evaluation of atopic status was not performed. 

6.13 SUMMARY 

Responses in humans exposed to ambient O3 concentrations include: decreased inspiratory 

capacity; mild bronchoconstriction; rapid, shallow breathing pattern during exercise; and 

symptoms of cough and pain on deep inspiration. Reflex inhibition of inspiration results in a 

decrease in forced vital capacity (FVC) and, in combination with mild bronchoconstriction, 

6-44
contributes to a decrease in the forced expiratory volume in 1 s (FEV1). In addition to 

physiological pulmonary responses and symptoms of breathing discomfort, O3 exposure also 

results in airway hyperresponsiveness, inflammation, immune system activation, and epithelial 

injury. With repeated O3 exposures over several days, spirometric and symptom responses 

become attenuated, but this tolerance is lost after about a week without exposure. Airway 

responsiveness also appears to be attenuated with repeated O3 exposures, but less so than FEV1. 

Unlike spirometric and symptom responses, airway inflammation and small airways dysfunction 

may not become attenuated by repeated O3 exposures. 

Healthy young adults exposed to O3 concentrations $0.08 ppm develop significant 

reversible, transient decrements in pulmonary function if minute ventilation ( V0 E ) or duration of 

exposure are increased sufficiently. The pattern of FEV1 response appears to depend on the O3 

exposure profile. Triangular exposure profiles can potentially lead to greater FEV1 responses 

than square wave exposures at equivalent average O3 doses. O3-induced decrements in FEV1 do 

not appear to depend on gender, race, body surface area, height, lung size, or baseline FVC in 

healthy young adults. Healthy children experience similar spirometric responses but lesser 

symptoms from O3 exposure relative to young adults. On average, spirometric and symptom 

responses to O3 exposure appear to decline with increasing age beyond about 18 years of age. 

There is a large degree of intersubject variability in physiologic and symptomatic responses of 

heathy adults exposed to O3. However, responses tend to be reproducible within a given 

individual over a period of several months. With increasing O3 concentration, the distribution of 

FEV1 decrements becomes asymmetrical, with a few individuals experiencing large decrements. 

There is a tendency for slightly increased spirometric responses in mild asthmatics and 

allergic rhinitics relative to healthy young adults. Spirometric responses in asthmatics appear to 

be affected by baseline lung function, i.e., responses increase with disease severity. With 

repeated daily O3 exposures, spirometric responses of asthmatics become attenuated; however, 

airway responsiveness becomes increased in subjects with preexisting allergic airway disease 

(with or without asthma). Possibly due to patient age, O3 exposure does not appear to cause 

significant pulmonary function impairment or evidence of cardiovascular strain in patients with 

cardiovascular disease or chronic obstructive pulmonary disease relative to healthy subjects. 

Available information on recovery from O3 exposure indicates that an initial phase of 

recovery in healthy individuals proceeds relatively rapidly, with acute spirometric and symptom 

6-45
responses resolving within about 2 to 4 h. Small residual lung function effects are almost 

completely resolved within 24 h. Effects of O3 on the small airways, assessed by persistent 

decrement in FEF25-75 and altered ventilation distribution, may be due in part to inflammation. 

Indeed, a prolonged recovery of residual spirometric decrements following the initial rapid 

recovery could be due to slowly resolving airway inflammation. In hyperresponsive individuals, 

this recovery takes longer (as much as 48 hours) to return to baseline values. Persistent 

spirometry changes observed for up to 48 h postexposure could plausibly be sustained by the 

inflammatory mediators. Cellular responses (e.g., release of immunomodulatory cytokines) 

appear to still be active as late as 20 h postexposure. More slowly developing inflammatory and 

cellular changes may persist for up to 48 h, but the time course for these parameters in humans 

has not been explored fully. 

Soluble mediators of inflammation such as the cytokines (IL-6, IL-8) and arachidonic acid 

metabolites (e.g., PGE2, PGF2", thromboxane, and leukotrienes [LTs] such as LTB4) have been 

measured in the BAL fluid of humans exposed to O3. Many of these compounds have 

bronchoconstrictive properties and may be involved in increased airway responsiveness 

following O3 exposure. Some indicators of inflammation (e.g., PMN influx, IL-6, PGE2, 

fibronectin) are attenuated with repeated O3 exposures. However, other markers (LDH, IL-8, 

total protein, epithelial cells) do not show attenuation, thus indicating that tissue damage 

probably continues to occur during repeated O3 exposure. There appears to be no strong 

correlation between any of the measured cellular and biochemical changes and changes in lung 

function measurements. A limited number of studies suggest that inflammatory responses may 

be detected following O3 exposures that are insufficient to cause decrements in pulmonary 

function. Whether airway reactivity or inflammatory responses to O3 are dependent on the age 

of the exposed individual, such as spirometric responses, has not been determined. 

Dietary antioxidant supplementation attenuates O3-induced spirometric responses, but not 

the intensity of subjective symptoms nor inflammatory responses. Dietary antioxidants also 

appear to afford partial protection to asthmatics by attenuating postexposure bronchial 

hyperresponsiveness. 

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7. EPIDEMIOLOGIC STUDIES OF

HUMAN HEALTH EFFECTS ASSOCIATED

WITH AMBIENT OZONE EXPOSURE

7.1 INTRODUCTION 

This chapter evaluates current epidemiologic literature on health and physiological effects 

of ambient ozone (O3) exposure. Epidemiologic studies linking community ambient O3 

concentrations to health effects were reported in the 1996 Ozone Air Quality Criteria Document 

(O3 AQCD; U.S. Environmental Protection Agency, 1996a). Many of those studies reported that 

pulmonary function decrements, respiratory symptoms, and hospital and emergency department 

admissions in human populations were associated with ambient levels of O3. Numerous more 

recent epidemiologic studies discussed in this chapter evaluate the relationship of ambient O3 to 

morbidity and mortality, and thereby provide an expanded basis for assessment of health effects 

associated with exposures to O3 at concentrations currently encountered in the United States. 

As discussed elsewhere in this document (Chapters 5 and 6), a substantial amount of 

experimental evidence links O3 exposure unequivocally with respiratory effects in laboratory 

animals and humans. These include structural changes in the bronchiolar-alveolar transition 

(centriacinar) region of the lung, biochemical evidence of acute cellular and tissue injury, 

inflammation, increased frequency and severity of experimental bacterial infection, and 

temporary reductions in mechanical lung function. These effects have been observed with 

exposure to O3 at ambient or near-ambient concentrations. Thus, many of the reported 

epidemiologic associations of ambient O3 with respiratory health effects have considerable 

biological credibility. Accordingly, the new epidemiologic studies of ambient O3 assessed here 

are best considered in combination with information from the other chapters on ambient O3 

concentration and exposure (Chapter 3), and toxicological effects of O3 in animals and humans 

(Chapters 5 and 6, respectively). The epidemiologic studies constitute important information on 

associations between health effects and exposures of human populations to “real-world” O3 and 

also help to identify susceptible subgroups and associated risk factors. 

7-1
A wide variety of oxidants in both the gaseous and particulate phases have not been 

examined in relation to health outcomes in the literature. Therefore, discussion in this chapter is 

limited to studies of human health effects associated with ambient O3 exposure. Ambient 

concentrations of the most abundant non-O3 oxidants (i.e., peroxyacetyl nitrate, peroxypropionyl 

nitrate, and H2O2) have not been shown to cause adverse health effects in toxicologic studies. 

However, as constituents of ambient air mixes, these oxidants may contribute to some of the 

effects attributed to O3. Therefore, health effect associations observed in relation to ambient O3 

concentrations may represent O3 effects, per se, or O3 may be serving as a surrogate measure for 

the overall photochemical oxidant mix. 

7.1.1 Approach to Identifying Ozone Epidemiologic Studies 

Numerous O3 epidemiologic papers have been published since completion of the 1996 O3 

AQCD. The U.S. Environmental Protection Agency (NCEA-RTP) has implemented a 

systematic approach to identify relevant epidemiologic studies for consideration in this chapter. 

In general, an ongoing search has been employed in conjunction with other strategies to 

identify O3 epidemiologic literature pertinent to developing criteria for the O3 National Ambient 

Air Quality Standards (NAAQS). A publication base was established using Medline, Pascal, 

BIOSIS, NTIS, and Embase, as well as a set of search terms proven by prior use to identify 

pertinent literature. The search strategy was later reexamined and modified to enhance 

identification of published papers. PubMed was added to the search regime. New studies 

accepted for publication through December 2004, as identified using the approaches above, have 

been included in this AQCD. 

While the above search regime provided good coverage of the relevant literature, 

additional approaches augmented the traditional search methods. First, a Federal Register 

Notice was issued requesting information and published papers from the public at large. Next, 

non-EPA chapter authors, expert in this field, identified literature on their own. NCEA-RTP 

staff also identified publications as an element of their assessment and interpretation of the 

literature. Finally, additional potentially relevant publications were included following external 

review as a result of comments from both the public and the Clean Air Scientific Advisory 

Committee (CASAC). More recent studies accepted in 2005 and 2006 for publication also were 

included if they added significantly to the existing body of data on critically important topics, 

7-2
such as those discussing methodological issues (e.g., O3 exposure misclassification, potential 

confounding by copollutants and meteorological factors), the O3-mortality relationship, and 

cardiovascular health outcomes. The combination of these approaches is believed to have 

produced a comprehensive ascertainment of studies appropriate for review and assessment here. 

The main criterion for selecting literature for the present assessment is to include those identified 

studies that evaluated relationships between measured ambient O3 levels and human health 

outcomes, which are pertinent to the evaluation of scientific bases useful for derivation of O3 

NAAQS for the United States. 

7.1.2 Approach to Assessing Epidemiologic Evidence 

Definitions of the various types of epidemiologic studies assessed here were noted in an 

earlier particulate matter (PM) AQCD (U.S. Environmental Protection Agency, 1996b). Briefly, 

epidemiologic studies are generally divided into two groups, morbidity studies and mortality 

studies. Morbidity studies evaluate O3 effects on a wide range of health endpoints, including: 

changes in pulmonary function, respiratory symptoms, self-medication in asthmatics, and airway 

inflammation; changes in cardiovascular physiology/functions; and cardiopulmonary emergency 

department visits and hospital admissions. Mortality studies investigate O3 effects on total 

(nonaccidental) mortality and cause-specific mortality, providing evidence related to a clearly 

adverse endpoint. The epidemiologic strategies most commonly used in O3 health studies are 

cross-sectional studies, prospective cohort studies, ecologic studies, time-series semi-ecologic 

studies, and case-crossover studies. All of these are observational rather than experimental 

studies. 

The approach to assessing epidemiologic evidence has been stated most recently in the 

2004 PM AQCD (U.S. Environmental Protection Agency, 2004) and is summarized here. The 

critical assessment of epidemiologic evidence presented in this chapter is conceptually based 

upon consideration of salient aspects of the evidence of associations so as to reach fundamental 

judgments as to the likely causal significance of the observed associations (see Hill, 1965). The 

general evaluation of the strength of the epidemiologic evidence reflects consideration not only 

of the magnitude and precision of reported O3 effect estimates and their statistical significance, 

but also of the robustness of the effects associations. Statistical significance corresponds to the 

allowable rate of error (Type I error) in the decision problem constructed from assuming that a 

7-3
simple null hypothesis of no association is true. It is a conditional probability; for statistical 

significance, typically there is a less than 0.05 chance of rejecting the null hypothesis given that 

it is true. Robustness of the associations is defined as stability in the effect estimates after 

considering a number of factors, including alternative models and model specifications, potential 

confounding by copollutants, as well as issues related to the consequences of measurement error. 

Consideration of the consistency of the effects associations, as discussed in the following 

sections, involves looking across the results of multiple- and single-city studies conducted by 

different investigators in different places and times. Relevant factors are known to exhibit much 

variation across studies, including, for example, the presence and levels of copollutants, the 

relationships between central measures of O3 and exposure-related factors, relevant demographic 

factors related to sensitive subpopulations, and climatic and meteorological conditions. Thus, in 

this case, consideration of consistency and the related heterogeneity of effects are appropriately 

understood as an evaluation of the similarity or general concordance of results, rather than an 

expectation of finding quantitative results within a very narrow range. 

Looking beyond the epidemiologic evidence, evaluation of the biological plausibility of 

the O3-health effects associations observed in epidemiologic studies reflects consideration of 

both exposure-related factors and dosimetric/toxicologic evidence relevant to identification of 

potential biological mechanisms. Similarly, coherence of health effects associations reported in 

the epidemiologic literature reflects consideration of information pertaining to the nature of the 

various respiratory- and cardiac-related mortality and morbidity effects and biological markers 

evaluated in toxicologic and human clinical studies. These broader aspects of the assessment are 

only touched upon here but are more fully integrated in the discussion presented in Chapter 8. 

In assessing the relative scientific quality of epidemiologic studies reviewed here and to 

assist in interpreting their findings, the following considerations were taken into account: 

(1) To what extent are the aerometric data/exposure metrics used of adequate quality and 

sufficiently representative to serve as credible exposure indicators, well-reflecting 

geographic or temporal differences in study population pollutant exposures in the 

range(s) of ambient pollutant concentrations evaluated? 

(2) Were the study populations well defined and adequately selected so as to allow for 

meaningful comparisons between study groups or for meaningful temporal analyses 

of health effects results? 

7-4
(3) Were the health endpoint measurements meaningful and reliable, including clear 

definition of diagnostic criteria utilized and consistency in obtaining dependent 

variable measurements? 

(4) Were the statistical analyses used appropriate, as well as properly performed and 

interpreted? 

(5) Were likely important covariates (e.g., potential confounders or effect modifiers) 

adequately controlled for or taken into account in the study design and statistical 

analyses? 

(6) Were the reported findings internally consistent, biologically plausible, and coherent 

in terms of consistency with other known facts? 

These guidelines provide benchmarks for judging the relative quality of various studies and 

in assessing the body of epidemiologic evidence. Detailed critical analysis of all epidemiologic 

studies on O3 health effects, especially in relation to all of the above questions, is beyond the 

scope of this document. As discussed in the upcoming Sections 7.1.3 and 7.1.4, considerations 

in the interpretation and presentation of the epidemiologic evidence led to emphasis being placed 

on certain studies in the main chapter text, tables, and figures. Additional studies in the O3 

epidemiologic literature are presented in Chapter 7 Annex Tables (see Annex Section AX7.1). 

Of most importance are those studies which provide useful qualitative or quantitative 

information on concentration-response relationships for health effects associated with ambient 

air levels of O3 likely to be encountered among healthy and susceptible populations in the 

United States. 

7.1.3 Considerations in the Interpretation of Epidemiologic Studies of 

Ozone Health Effects 

Prior to discussing results from recent O3 epidemiologic studies, issues and questions 

arising from the study designs and analysis methods used in the assessment of O3 effect 

estimates will be briefly presented. Study design can restrict the health effect parameters that 

can be estimated. Separate considerations need to be made for acute versus chronic effect 

studies, as well as individual- versus aggregate-level analyses. Time-series studies and panel 

studies are most frequently conducted in air pollution epidemiologic research. Aggregate-level 

exposure and/or outcome data are often used in these types of studies. Analyses using 

administrative health outcome data (e.g., numbers of deaths and emergency hospital admissions) 

7-5
have inherent limitations as well as strengths (Virnig and McBean, 2001). The impact of study 

design or the loss of information due to aggregation depends on the source of exposure 

(Sheppard et al., 2005). 

This section mainly focuses on the topics of exposure assessment and model specification 

in air pollution epidemiologic studies. Potential biases that may result from O3 exposure 

measurement error and from the choice of exposure index and lag period are discussed first. 

Model specification issues and potential confounding by temporal factors, meteorological 

effects, seasonal trends, and copollutants are then discussed. 

7.1.3.1 Exposure Assessment and Measurement Error in Epidemiologic Studies 

In many air pollution epidemiologic studies, especially time-series studies with 

administrative data on mortality and hospitalization outcomes, data from central ambient 

monitoring sites generally are used as the estimate of exposure. Personal exposures of individual 

study participants generally are not directly measured in epidemiologic studies. The use of O3 

concentrations from ambient monitors as surrogate measures for personal O3 exposures was 

discussed previously (Chapter 3, Section 3.9). Routinely collected ambient monitor data, though 

readily available and convenient, may not represent true personal exposure, which includes both 

ambient and non-ambient (i.e., indoor) source exposures. 

In several studies focused on evaluating exposure to O3, measurements were made in a 

variety of indoor environments, including homes (Lee et al., 2004), schools (Linn et al., 1996), 

and the workplace (Liu et al., 1995). Indoor O3 concentrations were, in general, approximately 

one-tenth of the outdoor concentrations in these studies. Few indoor sources of O3 exist, 

possible sources being office equipment (e.g., photocopiers, laser printers) and air cleaners. 

As described in Section 3.8 of this document, O3 in the indoor environment is largely dependent 

on the outdoor ambient O3 concentration. Other factors that influence indoor O3 concentrations 

include air exchange rate, outdoor infiltration, indoor circulation rate, and O3 removal processes. 

Sheppard (2005) described the relationship between panel studies and time-series studies 

and discussed the use of personal exposure measurements versus ambient concentrations in these 

designs, with the main focus being on nonreactive pollutants (e.g., PM). She noted that nonambient exposures typically varied across individuals but were not likely to have strong temporal 

correlations. In contrast, ambient concentrations across individuals should be highly correlated, 

7-6
as they tend to vary over time similarly for everyone because of changes in source generation, 

weather, and season. The independence of ambient and non-ambient exposure sources has 

important implications for selection of study designs that are most effective for estimating health 

effects (Sheppard, 2005). A simulation study by Sheppard et al. (2005) examining nonreactive 

pollutants found that there was no noticeable difference between effect estimates using either 

total personal exposure or ambient concentration data when non-ambient source exposures were 

independent of ambient source exposures in time-series studies. In the case of O3, there are 

limited non-ambient sources; thus, the non-ambient source exposures are likely to be 

independent from ambient source exposures. However, unlike PM, O3 is a reactive pollutant. 

In applying these conclusions to O3, an additional assumption needs to be made, i.e., that its 

chemical reactivity does not introduce strong temporal correlations. 

Other complications for O3 in the relationship between personal exposures and ambient 

concentrations include expected strong seasonal variation of personal behaviors and building 

ventilation practices that can modify exposure. In addition, the relationship may be affected by 

temperature (e.g., high temperature may increase air conditioning use, which may reduce O3 

penetration indoors), further complicating the role of temperature as a confounder of O3 health 

effects. It should be noted that the pattern of exposure misclassification error and influence of 

confounders may differ across the outcomes of interest as well as in susceptible populations. For 

example, those who may be suffering from chronic cardiovascular or respiratory conditions may 

be in a more protective environment (i.e., with less exposure to both O3 and its confounders, 

such as temperature and PM) than those who are healthy. 

As discussed thoroughly in the 2004 PM AQCD (Section 8.4.5), the resulting exposure 

measurement error and its effect on the estimates of relative risk must be considered. In theory, 

there are three components to exposure measurement error in time-series studies as described by 

Zeger et al. (2000): (1) the use of average population rather than individual exposure data; 

(2) the difference between average personal ambient exposure and ambient concentrations at 

central monitoring sites; and (3) the difference between true and measured ambient 

concentrations. The first error component, having aggregate rather than individual exposure 

data, is a Berksonian measurement error, which in a simple linear model increases the standard 

error, but does not bias the risk estimate. The second error component resulting from the 

difference between average personal ambient exposure and outdoor ambient concentration level 

7-7
has the greatest potential to introduce bias. If the error is of a fixed amount (i.e., absolute 

differences do not change with increasing concentrations), there is no bias. However, if the error 

is not a fixed difference, this error will likely attenuate the O3 risk estimate as personal O3 

exposures are generally lower than ambient O3 concentrations. The third error component, the 

instrument measurement error in the ambient levels, is referred to as nondifferential 

measurement error and is unlikely to cause substantial bias. 

The impact of exposure measurement error on O3 effect estimates was demonstrated in a 

study by Navidi et al. (1999). In this study, two statistical designs (bidirectional case-crossover 

design and multilevel analytic design) were considered. Simulations were conducted using data 

from the University of Southern California Children’s Health Study of the long-term effects of 

air pollutants on children. The effect estimate from computed “true” O3 exposure was compared 

to effect estimates from exposure determined by using several methods: (1) ambient stationary 

monitors; (2) the microenvironmental approach (multiply O3 concentrations in various 

microenvironments by time present in each microenvironment); and (3) personal sampling. 

Effect estimates based on all three exposure measures were biased toward the null. The bias that 

results when using the microenvironmental and personal sampling approach is due to 

nondifferential measurement error. Use of ambient monitors to determine exposure will 

generally overestimate true personal O3 exposure (because their use implies that subjects are 

outdoors 100% of their time and not in close proximity to sources that reduce O3 levels such as 

NO emissions from mobile sources); thus, generally, their use can result in effect estimates that 

are biased toward the null if the error is not of a fixed amount. 

Zidek (1997) noted that a statistical analysis must balance bias and imprecision (error 

variance). Ignoring measurement error in air pollution epidemiologic studies often results in 

underestimated risk estimates. For example, in a reanalysis of the study by Burnett et al. (1994) 

on the acute respiratory effects of ambient air pollution, Zidek et al. (1998) noted that accounting 

for measurement error, as well as making a few additional changes to the analysis, resulted in 

qualitatively similar conclusions. However, while the original analysis by Burnett et al. 

observed a 4.0% (95% CI: 1.9, 6.0) excess risk of daily respiratory admissions in the summer 

months attributable to a 40 ppb increase in 1-h max O3 (combined effect of a 1- and 3-day lag 

of O3), results from Zidek et al. indicated that a 9.4% (95% CI: 7.0, 11.8) increase was observed 

for a 40 ppb increase (a change from 10 to 50 ppb) in 1-h max O3 (combined effect of a 2- and 

7-8
3-day lag of O3). Although the different lag periods examined in the two analyses make direct 

comparisons of the effect estimates difficult, it appears that correcting the measurement error 

resulted in a larger effect estimate. Available data and analysis limit our ability to weigh the 

importance of uncertainty due to measurement error relative to other sources in the various 

studies reviewed in this chapter. 

In addition to overestimation of exposure and the resulting underestimation of effects, the 

use of ambient O3 concentrations may obscure the presence of thresholds in epidemiologic 

studies at the population level. Using PM2.5 as an example, Brauer et al. (2002) examined the 

relationship between ambient concentrations and mortality risk in a simulated population with 

specified common individual threshold levels. They found that no population threshold was 

detectable when a low threshold level was specified. Even at high specified individual threshold 

levels, the apparent threshold at the population level was much lower than specified. Brauer 

et al. (2002) concluded that surrogate measures of exposure (i.e., those from centrally-located 

ambient monitors) that were not highly correlated with personal exposures obscured the presence 

of thresholds in epidemiologic studies at the population level, even if a common threshold exists 

for individuals within the population. 

As discussed in Section 3.9, O3 concentrations measured at central ambient monitors may 

explain, at least partially, the variance of individual personal exposures; however, this 

relationship is influenced by factors such as air exchange rates in housing and time spent 

outdoors, which may vary by city. Other studies conducted in various cities observed that the 

daily averaged personal O3 exposures from the population were well correlated with monitored 

ambient O3 concentrations, although substantial variability existed among the personal 

measurements. Thus, there is supportive evidence that ambient O3 concentrations from central 

monitors may serve as valid surrogate measures for mean personal O3 exposures experienced by 

the population, which is of most relevance to time-series studies. This is especially likely true 

for respiratory hospital admission studies for which much of the response is attributable to O3 

effects on asthmatics. In children, for whom asthma is more prevalent, ambient monitors are 

likely to correlate reasonably well with personal exposure to O3 of ambient origin because 

children spend more time outdoors in the warm season. However, of some concern for mortality 

and hospitalization time-series studies is the extent to which ambient O3 concentrations are 

representative of personal O3 exposures in another particularly susceptible group of individuals, 

7-9
the debilitated elderly, as the correlation between the two measurements has not been examined 

in this population. A better understanding of the relationship between ambient concentrations 

and personal exposures, as well as of the factors that affect the relationship will improve the 

interpretation of ambient concentration-population health response associations observed. 

Existing epidemiologic models may not fully take into consideration all the biologically 

relevant exposure history or reflect the complexities of all the underlying biological processes. 

Using ambient concentrations to determine exposure generally overestimates true personal O3 

exposures (by approximately 2- to 4- fold in the various studies described in Section 3.9), 

resulting in biased descriptions of underlying concentration-response relationships (i.e., in 

attenuated risk estimates). The implication is that the effects being estimated occur at fairly low 

exposures and the potency of O3 is greater than these effect estimates indicate. As very few 

studies evaluating O3 health effects with personal O3 exposure measurements exist in the 

literature, effect estimates determined from ambient O3 concentrations must be evaluated and 

used with caution to assess the health risks of O3. 

The ultimate goal of the O3 NAAQS is to set a standard for the ambient level, not personal 

exposure level, of O3. Until more data on personal O3 exposure become available, the use of 

routinely monitored ambient O3 concentrations as a surrogate for personal exposures is not 

generally expected to change the principal conclusions from O3 epidemiologic studies. 

Therefore, population health risk estimates derived using ambient O3 levels from currently 

available observational studies (with appropriate caveats taking into account personal exposure 

considerations) remain useful. 

7.1.3.2 Ozone Exposure Indices Used 

The O3-related effect estimates for mortality and morbidity health outcomes are usually 

presented in this document as a relative risk, i.e., the risk rate relative to a baseline mortality or 

morbidity rate. Relative risks are based on an incremental change in exposure. To enhance 

comparability between studies, presenting these relative risks by a uniform exposure increment 

is needed. However, determining a standard increment is complicated by the use of different O3 

exposure indices in the existing health studies. The three daily O3 exposure indices that most 

often appear in the literature are the maximum 1-h average within a 24-h period (1-h max), the 

maximum 8-h average within a 24-h period (8-h max), and 24-h average (24-h avg) 

7-10
concentrations. As levels are lower and less variable for the longer averaging times, relative 

risks of adverse health outcomes for a specific numeric concentration range are not directly 

comparable across metrics. Using the nationwide distributional data for O3 monitors in U.S. 

Metropolitan Statistical Areas, increments representative of a low-to-high change in O3 

concentrations were approximated on the basis of annual mean to 95th percentile differences 

(Langstaff, 2003), as follows: 

Daily Exposure Index Exposure Increment (ppb) 

1-h max O3 40 

8-h max O3 30 

24-h avg O3 20 

In the following chapter sections, efforts were made to standardize the O3 risk estimates using 

these increments, except as noted. The specified incremental change for each daily O3 exposure 

index ensures that risk estimates are comparable across the different metrics. For example, a 2% 

excess risk for a 20 ppb change in mean 1-h max O3 was standardized and presented as a 4% 

excess risk for a 40 ppb change in mean 1-h max O3. This standardized risk estimate is 

approximately comparable to the excess risks observed for a 30 ppb change in mean 8-h max O3 

or a 20 ppb change in mean 24-h avg O3. Thus, the different increments for each daily O3 

exposure index do not represent inconsistencies; rather, they are appropriately scaled to facilitate 

comparisons between the various studies that used different indices. Note that in the Chapter 7 

Annex Tables (see Annex Section AX7.1), effect estimates are not standardized; there, the 

results are presented in the tables as reported in the published papers. 

7.1.3.3 Lag Time: Period between Ozone Exposure and Observed Health Effect 

Exposure lags may reflect the distribution of effects across time in a population and the 

potential mechanisms of effects. The choice of lag days for the relationship between exposure 

and health effects depends on the hypothesis being tested and the mechanism involved in the 

expression of the outcome. Effects can occur acutely with exposure on the same or previous 

day, cumulatively over several days, or after a delayed period of a few days. With knowledge 

of the mechanism of effect, the choice of lag days can be determined prior to analysis. 

7-11
As one example, one could expect cough to occur acutely after exposure with a lag of 0 or 1 day, 

given that O3 can act as a short-term irritant. However, an O3-related inflammatory response 

may not lead to asthma exacerbation until several days later. An asthmatic may be impacted 

by O3 on the first day of exposure, have further effects triggered on the second day, and then 

report to the emergency room for an asthmatic attack three days after exposure. Further, within 

a population of asthmatics, exacerbation of asthma symptoms may be observed over a period of 

several days, since each asthmatic may have varying individual aspects of the disease and may 

be affected by the exposure differently depending on his/her sensitivity and disease severity. 

The results from controlled human studies may be useful in assessing the adequacy of lags for 

some respiratory health outcomes. 

Some studies attempted to examine the overall impact of O3 through distributed lag 

models. The single-day lag model calculates a risk estimate that assumes dependence only on 

exposure from the specified day. In contrast, the distributed lag model provides an estimate that 

is a summary measure of the cumulative distributed lag effect from all included lag days. The 

standard error of the cumulative sum of the individual distributed lag coefficients takes into 

consideration the variance-covariance of the multiple lags, and it is generally larger than the 

standard error of the single-day lag coefficient due to positive auto-correlation. Thus, if the 

underlying O3-health outcome relationship was a single-day effect, then modeling the 

relationship with a distributed lag model would make the estimate less significant. On the other 

hand, if the effect of O3 on health outcomes persisted over several days, then applying a singleday lag model would result in an underestimation of the multiday effects, although the singleday estimate may still reflect a portion of the multiday effect, to the extent that daily O3 levels 

are auto-correlated. Choosing a lag model requires balancing variance and bias, and the 

usefulness (i.e., ease of interpretation, comparability across studies) of the model, as well as 

knowledge about mechanisms underlying effects for a given health outcome. Sufficient 

information is not available to judge the adequacy of a given model for different outcomes. 

As the parameters estimated from single-day lag versus multiday lag models are not the 

same, interpretation and comparison of these results will be difficult. When comparing the 

impacts of these different models, the nuance of increments used in calculating the estimates is 

different depending on the model. For example, an excess percent mortality risk “per 20 ppb 

increase in 24-h avg O3” in a distributed lag model including lag 0- through 6-days tacitly means 

7-12
a 20 ppb increase in each of the 7 days. The difference in the exposure scenarios in the singleday versus multiday lag model (i.e., 20 ppb increase in one day versus several consecutive days) 

complicates a simple comparison of risk estimates from two different models using “the same 

increment.” Thus, it should be recognized that these are distinct models that summarize the 

underlying process differently. 

Only a limited number of studies have hypothesized a priori the lag structure to be 

examined. Most of the O3 time-series studies examined relatively small numbers of single-day 

lag models, typically lags of 0 through 3 days. Sheppard et al. (1999) noted that when 

considering single-day lag estimates it is important to consider the effect estimate in the context 

of the pattern of adjacent lags as these estimates contain information from the adjacent days 

owing to serial correlation of the pollutant series. In many cases, a pattern of positive 

associations across several lag days was reported. For the respiratory and cardiovascular 

outcomes investigated, the “most significant” lags were generally 0- or 1-day lags, suggesting 

that the majority of the single-day associations are immediate, not a random pattern in which 

associations can be observed on any of the lags examined with equal probabilities. For example, 

two recent meta-analyses of O3-mortality effects observed that the combined estimate from 

0-day lag models was larger than the estimate from longer lag days (Bell et al., 2005; Levy 

et al., 2005). 

Bias resulting from the selection of lags has not been examined specifically for O3 effects. 

However, the issue of lags has been investigated for PM and the results of this analysis are most 

likely of relevance for O3. Lumley and Sheppard (2000) performed a simulation study to 

examine model selection bias in air pollution epidemiology using PM2.5 as an example. 

Sheppard et al. (1999; reanalysis Sheppard, 2003) had investigated the association between 

asthma hospital admissions and ambient PM2.5 concentrations over an eight-year period in 

Seattle, WA. Note that the results from Lumley and Sheppard (2000) and Sheppard et al. (1999) 

were based on Generalized Additive Model (GAM) analyses using default convergence criteria 

(see Section 7.1.3.7). A negative control analysis, using simulated data with no association 

between PM exposure and the health outcome, and a positive control analysis, using a specified 

non-zero excess risk added to the simulation, were performed for comparison. The bias from 

selection of the best of seven lags (0 to 6 days) and residual seasonal confounding in the negative 

control analysis (median log relative risk of 0.0013) was approximately half the log relative risk 

7-13
estimated from the observed data (0.0027), after adjusting for season and temperature. In the 

positive control model (true log relative risk of 0.0083), the bias was small (median log relative 

risk of 0.0080). Results from these simulations indicate that bias from selection of lags may be 

small, but of the same magnitude as the estimated health impacts. 

Selection of lag periods should depend on the hypothesis of the study and the potential 

mechanism of the effect. When the mechanism of the health effect is unknown, investigating the 

association between outcome and exposure using cumulative distributed lag models may be 

informative. Analyzing a large number of lags and simply choosing the largest and most 

significant results may bias the air pollution risk estimates away from the null. Most studies 

have shown that O3 has a fairly consistent, immediate effect on health outcomes, including 

respiratory hospitalizations and mortality. Several studies also observed significant O3 effects 

over longer cumulative lag periods, suggesting that in addition to single-day lags, multiday lags 

should be investigated to fully capture a delayed O3 effect on health outcomes. In this document, 

discussion largely focuses on effect estimates from 0- and 1-day lags, with some consideration of 

cumulative, multiday lag effects. It is not straightforward to compare and contrast results from 

single-day versus multiday lag models, because the parameters estimated from these models are 

not the same. These complications need to be taken into consideration when interpreting results 

from various lag models. 

7.1.3.4 Model Specification to Adjust for Temporal Trends and Meteorologic Effects 

Several challenges are encountered with respect to designing and interpreting time-series 

studies. The principal challenge facing the analyst in the daily time-series context is avoiding 

bias due to confounding by short-term temporal factors operating over time scales from days to 

seasons. In the current regression models used to estimate short-term effects of air pollution, 

two major potential confounders generally need to be considered: (1) seasonal trend and other 

“long-wave” temporal trends; and (2) weather effects. Both of these variables tend to predict a 

significant fraction of fluctuations in time-series. Unfortunately, as O3 has strong seasonal 

cycles and is formed more at higher temperatures, both terms are also highly correlated with O3. 

The correlation of O3 with these confounding terms tends to be higher than that for PM or other 

gaseous pollutants. In the United States, the mass concentration of PM2.5 generally does not 

have strong seasonal cycles like O3, because PM2.5 tends to reflect both primary emissions 

7-14
(throughout the year, but often higher in winter in most U.S. cities) and secondary aerosols 

(higher in summer). Therefore, PM2.5 and O3 effect estimates from studies primarily designed to 

examine PM2.5 health effects may not be comparable, because model specifications that may be 

appropriate for PM2.5 may not necessarily be adequate for O3. 

An examination of recent time-series studies indicates that several types of fitting 

approaches have been used to adjust for temporal trends and weather effects. The use of 

parametric and nonparametric smoothers with varying degrees of freedom per year has emerged 

as the prevailing approach. The use of larger degrees of freedom to adjust for potential 

confounding by time-varying factors may inadvertently result in ascribing more effects to these 

unmeasured potential confounders and mask the air pollution effect. Often smaller pollution 

effect estimates are observed when more degrees of freedom are used. Currently, the degrees of 

freedom used to adjust for temporal trends in time-series studies generally range from 4 to 

12 degrees of freedom per year using either nonparametric or parametric smoothers. Statistical 

diagnostics such as Akaike’s Information Criteria, residual autocorrelation, or dispersion of the 

regression model often are used to choose or to evaluate the adequacy of the degrees of freedom 

for temporal trend. However, these diagnostics do not guarantee “adequate” control for temporal 

confounding, as the appropriate extent of smoothing is not identifiable from the data and the 

proper selection of smoothing parameters requires prior knowledge of the nature of the 

confounding (e.g., shape and duration of influenza epidemics). 

The issue of model specifications to adjust for temporal trends and weather variables in 

time-series studies was a consideration of several researchers that conducted sensitivity analyses 

of PM data (Health Effects Institute, 2003). The sensitivity of O3 coefficients to model 

specifications for temporal trend adjustment has not been as well-studied. Recent multicity 

studies examined the sensitivity of O3 coefficients to the extent of smoothing for adjustment of 

temporal trends and meteorologic factors (Bell et al., 2004; Huang et al., 2005; Ito et al., 2005). 

Most, if not all, O3 studies used the same model specifications to estimate the excess risks for 

PM and other gaseous pollutants. The model specification designed to control confounding by 

meteorological and temporal factors for PM may not necessarily be adequate for O3. As noted 

above, O3 is expected to have the strongest correlation with both temporal (seasonal) trend and 

weather effects. The strong annual cycle in O3 concentrations presents a unique problem in 

time-series analyses where time trends are fitted simultaneously with pollution and other model 

7-15
terms (i.e., co-adjustment). In this setting, the annual O3 cycle itself may compete with the 

smooth function of time to explain some of the annual, cyclic behavior in the health outcome, 

which can result in biased effect estimates for O3 when data for all seasons are analyzed 

together. 

Current weather models used in time-series analyses can be classified by their use of: 

(1) quantile (e.g., quartile, quintile) indicators; (2) parametric functional forms such as V- or 

U-shape functions; and (3) parametric (e.g., natural splines) or nonparametric (e.g., locally 

estimated smoothing splines [LOESS]) smoothing functions. More recent studies tend to use 

smoothing functions. While these methods provide flexible ways to fit health outcomes as a 

function of temperature and other weather variables, there are two major issues that need further 

examination to enable more meaningful interpretation of O3 morbidity and mortality effects. 

The first issue is the interpretation of weather or temperature effects. Most researchers 

agree about the morbidity and mortality effects of extreme temperatures (i.e., heat waves or cold 

spells). However, as extreme hot or cold temperatures, by definition, happen rarely, much of the 

health effects occur in the mild or moderate temperature range. Given the significant correlation 

between O3 and temperature, ascribing the association between temperature and health outcomes 

solely to temperature effects may underestimate the effect of O3. The second issue is that 

weather model specifications are fitted for year-round data in most studies. Such models will 

ignore the correlation structure that can change across seasons, resulting in inefficiency 

and model mis-specification. This is particularly important for O3, which appears to change 

its relationship with temperature as well as with other pollutants across seasons. 

This changing relationship between O3 and temperature, as well as between O3 and other 

pollutants across seasons, and its potential implications for health effects modeling have not 

been examined thoroughly in the time-series literature. Even the flexible smoother-based 

adjustments for seasonal and other time-varying variables cannot fully take into account these 

complex relationships. One obvious way to alleviate or avoid this complication is to analyze 

data by season. While this practice reduces sample size, its extent would not be as serious as for 

PM (which is collected only every sixth day in most locations) because O3 is collected daily, 

though only in warm seasons in some states. An alternative approach is to include separate O3 

concentration variables for each season (by multiplying O3 concentrations by a season indicator 

7-16
variable). However, this approach assumes that all effects in the model that are not indicated to 

be season-specific do not vary seasonally. 

In locations where seasonal variability may be a factor, O3 effect estimates calculated using 

year-round data can be misleading, as the changing relationship between O3, temperature, and 

other pollutants across seasons may have a significant influence on the estimates. Analyses have 

indicated that confounding from seasonal variability may be controlled effectively by stratifying 

the data by season. 

7.1.3.5 Confounding Effects of Copollutants 

Extensive discussion of issues related to confounding effects among air pollutants in timeseries studies are provided in Section 8.4.3 of the 2004 PM AQCD (U.S. Environmental 

Protection Agency, 1996b). Since the general issues discussed there are applicable to all 

pollutants, such discussions are not repeated here. What was not discussed in the 2004 PM 

AQCD was the issue of changing relationships among air pollutants across seasons. Compared 

to other pollutants, O3 has strong seasonal cycles. Ambient O3 levels are typically higher in the 

summer or warm season, often referred to as the O3 season. In the winter or colder months, O3 

levels tend to be much lower than in the summer months. During the winter in some urban 

locations, O3 mainly comes from the free troposphere and can be considered a tracer for 

relatively clean air (i.e., cold, clear air coming down from the upper atmosphere), as discussed in 

Chapter 3 of this document. The clean air is associated with the passage of cold fronts and the 

onset of high-pressure conditions, which occur with colder temperatures. Thus, sunny clear 

winter days following a high-pressure system are the days when air pollution levels from 

primary emissions (e.g., NO2, SO2, and PM from local sources) tend to be lower and O3 is 

relatively higher. This can lead to negative correlations between O3 and the primary pollutants 

in the winter. As shown in Figure 3-21 in Chapter 3, the relationship between O3 and PM2.5 was 

U-shaped for the year-round data in Fort Meade, MD. The negative PM2.5/O3 slope was in the 

range of O3 concentrations less than 30 ppb, providing supporting evidence of the 

aforementioned winter phenomenon. Thus, the correlation between O3 and PM for year-round 

data may be misleading. The high reactivity of O3 with certain copollutants further complicates 

the analysis and interpretation. For example, the reaction between NO (emitted from motor 

vehicles) and O3 results in reduced O3 levels but increased NO2 levels during high traffic periods. 

7-17
Multipollutant regression models often are used to assess potential confounding by 

copollutants; however, there are limitations to these models. Zidek et al. (1996) examined, 

through simulation, the joint effects of multicollinearity and measurement error in a Poisson 

regression model. The results illustrated the transfer of effects from the “causal” variable to the 

confounder. However, in order for the confounder to have a larger effect size than the true 

predictor, the correlation between the two covariates had to be very high (r $ 0.9), with moderate 

error (" > 0.5, where " is the probability of rejecting the null hypothesis given that the null is 

true) for the true predictor and no error for the confounder in their scenarios. The transfer-ofcausality effect was lessened when the confounder also became subject to error. Another 

interesting finding was the behavior of the standard errors of the coefficients. When the 

correlation between the covariates was high (r = 0.9) and both covariates had no error, the 

standard errors for both coefficients were inflated by a factor of two; however, this phenomenon 

disappeared when the confounder had error. The effect of multicollinearity is generally even 

more complex when analyzing real data. For further discussion, see the 2004 PM AQCD 

(Sections 8.4.3 and 8.4.5). 

Uncertainty remains as to the use of multipollutant regression models in assessing the 

independent health effects of pollutants that are correlated. Particularly in the case of O3, 

concern remains as to whether multipollutant regression models for year-round data can adjust 

for potential confounding adequately because of the changing relationship between O3 and other 

pollutants. Despite these limitations, use of multipollutant models is still the prevailing 

approach employed in most, if not all, studies of O3 health effects; and it serves as an important 

tool in addressing the issue of confounding by copollutants, especially in season-stratified 

analyses. 

7.1.3.6 Hypothesis Testing and Model Selection in Ozone Epidemiologic Studies 

Epidemiologic studies investigated the association between various measures of O3 (e.g., 

multiple lags, different metrics, etc.) and various health outcomes using different model 

specifications. Statistically testing a null hypothesis (i.e., there is no effect of O3) requires one to 

calculate the value of a test statistic (i.e., the t-value). If the observed test statistic exceeds a 

critical value (oftentimes the 95th percentile) or is outside a range of values, the null hypothesis 

is rejected. However, when multiple testing is done using a critical value determined for a single 

7-18
test, the chance that at least one of the hypotheses is significant may be greater than the specified 

error rate. Procedures are available to ensure that the rejection error rate does not exceed the 

expected error rate (usually 5%) when conducting multiple hypothesis testing. However, often 

the multiple hypotheses being tested are not statistically independent, thus some corrections, 

such as the Bonferroni adjustment, may be overly conservative. 

Multiple hypothesis testing and model selection also contribute to publication bias. 

Publication bias is the tendency of investigators to submit and/or editors to accept manuscripts 

for publication based on the strength of the study findings. Although publication bias commonly 

exists for many topics of research, it may be present to a lesser degree in the air pollution 

literature. Several air pollutants often are examined in a single study, leading to the publication 

of significant, as well as nonsignificant, individual pollutant results. For example, many air 

pollution papers with a focus on PM health effects also published O3 results. Ozone was largely 

considered a potentially confounding copollutant of PM; thus, O3 effect estimates were often 

presented regardless of the statistical significance of the results. Another aspect of publication 

bias is only selecting the largest or statistically strongest effect estimate to report and not the 

array of models evaluated. Bell et al. (2005) conducted a meta-analysis of O3 health effects 

studies and found that studies reporting a single estimate compared to multiple estimates 

generally showed larger effects for the same lag. They also examined the presence of this bias 

by comparing the results from the meta-analysis to the multicity National Morbidity, Mortality, 

and Air Pollution Study (NMMAPS). When comparing the marginal posterior distributions of 

the overall effect under the meta-analysis and in NMMAPS for the eight U.S. cities common to 

both the approaches, the meta-analysis effect estimate of same-day O3 concentration was nearly 

2-fold that of the NMMAPS pooled effect estimate. 

Testing multiple hypotheses may, at times, be appropriate. For example, without an a 

priori biologic model that definitively specifies which of several lag periods or meteorologiccontrol models should take priority, multiple hypotheses may need to be developed for 

researchers to explore more thoroughly potential associations for an O3-related health effect 

(Goodman, 2005). In this case, it would be useful to state which hypotheses are confirmatory 

versus exploratory. The key concern is the reporting of a single estimate or estimates from one 

model. Presentation of an array of models, exposure definitions, and outcome measures may be 

more appropriate in communicating the results. Sensitivity analyses, which are critical for 

7-19
model validation, also involve multiple testing. When conducting sensitivity analyses, one 

should guard against multiple testing errors by restricting the inferences to consistency of 

the effect. 

PM

One method to address model uncertainty and multiple hypothesis testing is Bayesian 

model averaging. In Bayesian model averaging, predictions and inferences are based on a set of 

models rather than a single model, and each model contributes proportionally to the support it 

receives from the observed data (Clyde, 1999). In addition to the uncertainty of effect 

estimation, Bayesian model averaging can incorporate uncertainty regarding the choice of 

confounding variables, pollutants, and lags. Koop and Tole (2004) used Bayesian model 

averaging to analyze the effect of various air pollutants, including O3, SO2, CO, NO, NO2, 

10-2.5, and PM2.5, on mortality in Toronto, Canada. The 50+ explanatory variables required the 

fitting of an enormous number of potential models. Although the point estimates for all 

pollutants were positive, very small effects were found. However, in the context of the many 

interaction terms, meteorological variables, smoothing surfaces, and the relatively loose 

posterior distribution, the analysis results cannot be interpreted meaningfully. Clyde (2000) and 

Clyde et al. (2000) also used Bayesian model averaging to analyze the relationship between PM 

and mortality. Clyde (2000) noted that Bayesian model averaging did not take into consideration 

factors that might bias the estimated effect toward the null. For example, measurement error in 

the exposure variables was not considered. In addition, the Poisson model (similar to many 

other regression models) assumed that all individuals in a population had equal risks, including 

potentially susceptible and vulnerable populations such as those with respiratory illnesses and 

outdoor workers. While Bayesian model averaging can theoretically be used to take into account 

uncertainty, claims of causality based on observational studies may be highly sensitive to the 

choice of prior distributions and class of models under consideration (Clyde et al., 2000). 

Another limitation of Bayesian model averaging is that the estimated posterior effects may be 

diluted (i.e., result in smaller coefficients) when variables are highly correlated, as may be the 

case for air pollution studies (George, 1999 in comments to Hoeting et al., 1999). 

Additional methods to control for multiple hypothesis testing are by deciding a priori 

which hypotheses are confirmatory and exploratory, and then limiting the number of 

confirmatory tests. For example, Dominici et al. (2003) used a minimum number of tests in the 

U.S. 90 cities study, which reduced the uncertainty associated with multiple testing. In addition, 

7-20
they performed sensitivity analyses to examine the consistency and robustness of the effects. 

Another approach is to partition the data into two sets, one for model identification and a second 

for model confirmation. 

The summary of health effects in this chapter is vulnerable to the errors of publication bias 

and multiple testing. Efforts have been made to reduce the impact of multiple testing errors on 

the conclusions in this document. To address multiple hypothesis testing, emphasis will be 

placed in this chapter on a priori hypotheses. As identifying a priori hypotheses is difficult in the 

majority of the studies, the most common hypotheses will be considered. For example, although 

many studies examined multiple single-day lag models, priority would be given to effects 

observed at 0- or 1-day lags rather than at longer lags. Both single- and multiple-pollutant 

models that include O3 will be considered and examined for robustness of results. Analyses of 

multiple model specifications for adjustment of temporal or meteorological trends will be 

considered sensitivity analyses. Sensitivity analyses shall not be granted the same inferential 

weight as the original hypothesis-driven analysis; however, these analyses will be discussed in 

this chapter as appropriate given their valuable insights that may lead scientific knowledge in 

new directions. 

7.1.3.7 Impact of Generalized Additive Models Convergence Issue on Ozone Risk 

Estimates 

Generalized Additive Models (GAM) have been widely utilized for epidemiologic analysis 

of the health effects attributable to air pollution. The impact of the GAM convergence issue was 

thoroughly discussed in Section 8.4.2 of the 2004 PM AQCD. Reports have indicated that using 

the default convergence criteria in the Splus software package for the GAM function can lead to 

biased regression estimates for PM and an underestimation of the standard error of the effect 

estimate (Dominici et al., 2002; Ramsay et al., 2003). GAM default convergence criteria has a 

convergence precision of 10!3

 and a maximum number of 10 iterations. The more stringent 

convergence criteria refers to increased stringency of both the convergence precision and 

number of iterations. The use of default convergence criteria was found to be a problem when 

the estimated relative risks were small and two or more nonparametric smoothing curves were 

included in the GAM (Dominici et al., 2002). The magnitude and direction of the bias depend in 

part on the concurvity of the independent variables in the GAM and the magnitude of the risk 

7-21
estimate. Recent focus has been on the influence of the GAM function on effect estimates for 

PM. However, because O3 covaries more strongly with both weather and time factors than does 

PM, the issue of GAM convergence criteria for O3 also needs to be considered. 

A meta-analysis by Stieb et al. (2003) found some difference in O3-mortality risk estimates 

between the GAM and non-GAM studies. GAM studies were defined as studies that analyzed 

effect estimates using nonparametric smoothing functions of time or weather. Non-GAM studies 

were all other studies, including those using Generalized Linear Models (GLM) and Generalized 

Estimating Equations (GEE) in their analysis. In the single-pollutant models, the O3-mortality 

risk estimates for the non-GAM studies (10 estimates) and GAM studies (15 estimates) were 

1.8% (95% CI: 0.5, 3.1) and 2.2% (95% CI: 1.4, 2.8), respectively, per 40 ppb daily 1-h 

max O3. In the multipollutant models, the pooled risk estimate was 1.0% (95% CI: !0.5, 2.6) 

for non-GAM studies (7 estimates) and 0.5% (95% CI: !1.0, 1.9) for GAM studies 

(4 estimates). 

Results from recent meta-analyses of O3-mortality effects suggest that there are no 

substantial differences between GAM-affected estimates and non-GAM-affected estimates (Bell 

et al., 2005; Ito et al., 2005; Levy et al., 2005). GAM-affected studies included those that used 

default convergence criteria. Non-GAM-affected studies included GAM studies that used 

stringent convergence criteria or those that used other modeling techniques. Ito et al. (2005) 

found that the single-pollutant combined estimate for the GAM-affected studies (15 estimates) 

and non-GAM-affected studies (28 estimates) were 1.92% (95% CI: 1.02, 2.81) and 1.40% 

(95% CI: 0.78, 2.02), respectively, per 20 ppb increase in 24-h avg O3. In the analysis by Levy 

et al. (2005), the single-pollutant combined estimate for the GAM-affected studies (29 estimates) 

and non-GAM-affected studies (17 estimates) were 1.56% (95% CI: 1.01, 2.11) and 1.80% 

(95% CI: 1.17, 2.43), respectively, per 40 ppb increase in 1-h max O3. Bell et al. (2005) also 

reported that the pooled estimate was larger for the studies that were not GAM-affected. 

A few GAM studies reanalyzed O3 risk estimates using more stringent convergence criteria 

or GLM. Reanalysis of an asthma hospital admissions study in Seattle, WA (Sheppard et al., 

1999; reanalysis Sheppard, 2003) indicated that there were only slight changes in the risk 

estimates when using more stringent convergence precision (10!8

) in GAM. The original GAM 

analysis indicated an excess risk of 9% (95% CI: 3, 17), whereas the stringent GAM analysis 

found an excess risk of 11% (95% CI: 3, 19) per 30 ppb increase in 8-h max O3 at a 2-day lag. 

7-22
Similar results were found using GLM with natural splines: 11% (95% CI: 2, 20). In the 

reanalysis of Santa Clara County, CA data, Fairley (1999; reanalysis Fairley, 2003) used the 

same methods as the original analysis except the convergence precision (,) was increased 

from 10!4

 to 10!12 and the maximum number of iterations (M) were increased from 10 to 107. 

The O3-mortality risk estimate slightly increased from 2.8% (95% CI not provided) using default 

GAM parameters to 2.9% (95% CI: !0.3, 6.0) using stringent GAM parameters per 30 ppb 

increase in 8-h max O3 at a 0-day lag. The O3-mortality risk estimates further increased to 3.0% 

(95% CI: !0.3, 6.3) using GLM with natural cubic splines. In the reanalysis of the Netherlands 

data by Hoek et al. (2000; reanalysis Hoek, 2003), the O3 nonaccidental mortality risk estimates 

increased from 1.3% (95% CI: 0.8, 1.9) using default GAM to 1.5% (95% CI: 1.0, 2.1) using 

stringent GAM (, = 10!8

, M = 103) and 1.6% (95% CI: 0.9, 2.4) using GLM with natural splines 

per 30 ppb increase in 8-h avg O3 (12 p.m.-8 p.m.) at a 1-day lag. 

In the limited number of studies that have reanalyzed O3 risk estimates, there is little 

evidence that default GAM analyses resulted in positively biased estimates, as was observed for 

PM. Generally, it appears that the use of default convergence criteria in GAM tends to bias risk 

estimates toward the null, in addition to underestimating the standard errors. However, one 

study by Cifuentes et al. (2000) in Santiago, Chile observed a large difference in the O3­

mortality excess risks calculated using default GAM (0.9% [95% CI: 0.2, 1.6] per 40 ppb 

increase in 1-h max O3) and GLM (0.1% [95% CI: !0.6, 0.8]). The GAM convergence problem 

appears to vary depending on data sets, and likely depends upon the intercorrelation among 

covariates and the magnitude of the risk estimate; thus, its impact on the results of individual 

studies cannot be known without a reanalysis. Consistent with the approach used in the 2004 

PM AQCD, the results from studies that analyzed data using GAM with default convergence 

criteria and at least two nonparametric smoothing terms are generally not considered in this 

chapter, with some exceptions as noted. 

7.1.3.8 Summary of Considerations in the Interpretation of Ozone Epidemiologic Studies 

The previous sections discussed the topics of exposure assessment and model specification 

in O3 epidemiologic studies. Also examined were the issues of hypothesis testing and model 

selection, as well as the impact of the GAM convergence issue on O3 risk estimates. 

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C Exposure measurement error may result from the use of stationary ambient monitors 

as an indicator of personal exposure in population studies. The relationship between 

individual personal O3 exposures and ambient O3 concentrations was found to be 

affected by factors such as air exchange rates in housing and time spent outdoors, 

which varied by individual or by city. However, ambient O3 concentrations, though 

they do tend to overestimate personal exposures, were found in general to be well 

correlated to daily averaged personal O3 exposures for the populations studied. 

Therefore population risk estimates derived from ambient O3 concentrations remain 

useful and appropriate, but should be evaluated and used with caution in O3 health 

risk assessments. 

C The three daily O3 exposure indices used most often in O3 epidemiologic studies are 

1-h max O3, 8-h max O3, and 24-h avg O3. To facilitate comparison of health effects 

calculated using the different O3 metrics, risk estimates are standardized or scaled by 

specified exposure increments: 40 ppb for 1-h max O3, 30 ppb for 8-h max O3, and 

20 ppb for 24-h avg O3. 

C The lag period between O3 exposure and the observed health effect may reflect the 

distribution of effects across time in a population and the potential mechanisms of 

effects. Among the single-day lags examined, the strongest health effects were 

observed with a 0- or 1-day lag in O3 exposure. Multiday lags of O3 exposure also 

were investigated. As the parameters estimated from single-day versus multiday 

models are not the same, interpreting results from these distinct lag models 

requires caution. 

C In time-series studies estimating short-term effects of air pollution, temporal trends 

and weather effects are two major potential confounders that require consideration. 

Of all the air pollutants, O3 is expected to have the strongest correlation with both 

temporal and meteorological factors. The use of parametric and nonparametric 

smoothers with varying degrees of freedom per year has been the prevailing approach. 

Confounding from seasonal variability may be controlled effectively by stratifying 

analyses by season. 

C A major methodological issue affecting O3 epidemiologic studies concerns the 

evaluation of the extent to which other air pollutants may confound or modify 

O3-related effect estimates. The changing relationship between O3 and copollutants 

across seasons further complicates the issue. The use of multipollutant regression 

models is the prevailing approach for controlling potential confounding 

by copollutants in O3 health effects studies. 

C In air pollution epidemiologic studies, multiple hypotheses are often tested and certain 

models are selected for presentation. In many cases, the multiple hypotheses being 

tested area not independent and, thus, some correlations (e.g., the Bonferroni 

adjustment) may be overly conservative. To reduce the potential impact of multiple 

testing errors, emphasis should be placed on a limited number of a priori hypotheses. 

Additional analyses of varying model specifications considered along with sensitivity 

analyses. 

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C GAM is widely used in epidemiologic analyses of air pollution health effects. 

Recent reports have indicated that the commonly used default convergence criteria 

lead to biased regression estimates for PM. The limited number of studies that have 

evaluated this issue for O3 observed little evidence that default GAM analyses resulted 

in positively biased estimates. 

7.1.4 Approach to Presenting Ozone Epidemiologic Evidence 

To produce a thorough appraisal of newly available evidence, key information (including 

study design, analysis, mean O3 concentrations, and health outcome results) from important new 

studies is presented in summary tables in the Chapter 7 Annex (Annex Section AX7.1). Each 

section of this chapter starts by concisely highlighting important points derived from the 1996 O3 

AQCD assessment. In the main body of the chapter, particular emphasis is focused on studies 

and analyses that provide pertinent information for the critical assessment of health risks from O3 

exposure. Not all studies are accorded equal weight in the overall interpretive assessment of 

evidence regarding O3-associated health effects. Among well-conducted studies with adequate 

control for confounding, increasing scientific weight is accorded in proportion to the precision of 

their effect estimates. Small-scale studies without a wide range of exposures generally produce 

less precise estimates compared to larger studies with a broad exposure gradient. The size of the 

study, as indicated by the length of the study period and total number of events, and the 

variability of O3 exposures are important components that help to determine the precision of the 

health effect estimates. In evaluating the epidemiologic evidence in this chapter, more weight is 

accorded to estimates from studies with narrow confidence bands. 

Emphasis is placed in the text on the discussion of (1) new multicity studies that employ 

standardized methodological analyses for evaluating O3 effects across several or numerous cities 

and often provide overall effect estimates based on combined analyses of information pooled 

across multiple cities; (2) studies that consider O3 as a component of a complex mixture of air 

pollutants including PM and other gaseous criteria pollutants (CO, NO2, SO2); and (3) North 

American studies conducted in the United States or Canada. Multicity studies are of particular 

interest and value due to their evaluation of a wider range of O3 exposures and large numbers of 

observations. They generally provide more precise effect estimates than most smaller scale 

studies of single cities. Compared to meta-analyses of multiple “independent” studies, a 

potential advantage of multicity studies is their consistency in data handling and model 

7-25
specifications, which eliminates variation due to analysis approach. Also, unlike meta-analyses, 

they do not suffer from potential omission of nonsignificant results due to “publication bias.” 

Furthermore, geographic patterns of air pollution effects have the potential to provide especially 

valuable evidence regarding relative homogeneity and/or heterogeneity of O3 health effects 

relationships across geographic locations. Due to the potential for confounding by copollutants, 

preference is given to studies with effect estimates from multipollutant models, i.e., models with 

both O3 and PM rather than O3-only models. The potential impacts of different health care 

systems and the underlying health status of populations also need to be accounted for in the 

assessment (Hubbell et al., 2005; Levy et al., 2001); thus, U.S. studies are emphasized over 

non-U.S. studies. In accordance to the emphasis placed on the O3 epidemiologic studies in this 

chapter, the tables in the Chapter 7 Annex Section AX7.1 were organized by region, with 

multicity studies in each region presented first. 

In the coming sections, field/panel studies and studies of emergency department visits and 

hospital admissions, which contributed to the establishment of the revised 1997 NAAQS for O3, 

are presented first. This is followed by a discussion of O3-related mortality and effects of 

chronic exposures to O3. The chapter ends with an integrative discussion providing a summary 

and conclusions. 

7.2 FIELD STUDIES ADDRESSING ACUTE EFFECTS OF OZONE 

7.2.1 Summary of Key Findings on Field Studies of Acute Ozone Effects 

from the 1996 Ozone AQCD 

In the 1996 O3 AQCD, individual-level camp and exercise studies provided useful 

quantitative information on the concentration-response relationships linking human lung 

function declines with ambient O3 concentrations. The available body of evidence supported a 

dominant role of O3 in the observed lung function decrements. Extensive epidemiologic 

evidence of pulmonary function responses to ambient O3 has been derived from camp studies. 

Six studies from three separate research groups provided a combined database on individual 

concentration-response relationships for 616 children (mostly healthy, nonasthmatic) ranging in 

age from 7 to 17 years, each with at least six sequential measurements of FEV1 (forced 

expiratory volume in 1 second) while attending summer camps (Avol et al., 1990; Higgins et al., 

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1990; Raizenne et al., 1987, 1989; Spektor et al., 1988a, 1991). In the combined reanalysis by 

Kinney et al. (1996a) using consistent analytical methods, these data yielded an average 

relationship between afternoon FEV1 and concurrent-hour O3 concentration of !0.50 mL/ppb 

(95% CI: !0.63, !0.36), with study-specific slopes ranging from !1.29 to !0.19 mL/ppb. 

At the time of the afternoon lung function measurement, mean 1-h avg O3 concentrations ranged 

from 53 to 123 ppb (maximum range 95 to 245 ppb). Exposure in camp studies usually extended 

for multiple hours. Although the regression results noted above were based on 1-h O3 levels, 

single- and multiple-hour averages were observed to be highly correlated; thus, these results 

might represent, to some extent, the influence of multihour exposures. In addition to the camp 

study results, two studies involving lung function measurements before and after well-defined 

exercise events in adults yielded concentration-response slopes of !0.4 mL/ppb (95% CI: !0.7, 

!0.1) (Selwyn et al., 1985) and !1.35 mL/ppb (95% CI: !2.04, !0.66) (Spektor et al., 1988b). 

Ozone concentrations during exercise events of approximately 1⁄2-hour duration ranged from 4 to 

135 ppb in these studies. 

Results from other field/panel studies also supported a consistent relationship between 

ambient O3 exposure and acute respiratory morbidity in the population. Respiratory symptoms 

(or exacerbation of asthma) and decrements in peak expiratory flow (PEF) occurred with 

increased ambient O3 concentrations, especially in asthmatic children (Lebowitz et al., 1991; 

Krzyzanowski et al., 1992). The results showed greater responses in asthmatic than in 

nonasthmatic individuals (Lebowitz et al., 1991; Krzyzanowski et al., 1992), indicating that 

asthmatics might constitute a sensitive group in epidemiologic studies of oxidant air pollution. 

Since the 1996 O3 AQCD, new research has examined a broad scope of field studies as 

discussed next. 

7.2.2 Introduction to Recent Field Studies of Acute Ozone Effects 

Numerous field studies carried out over the past decade have tested for and, in many cases, 

observed acute associations between O3 concentrations and measures of respiratory ill-health and 

O3 concentrations in groups of subjects (Table AX7-1 in Chapter 7 Annex Section AX7.1). 

Acute field studies are distinguished from time-series studies in that they are designed to recruit 

and collect data from individual human subjects instead of utilizing administrative data on 

aggregate health outcomes such as daily mortality, hospital admissions, or emergency 

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department visits. Although individual-level health outcome data are collected in field studies, 

ambient O3 concentrations from centrally located monitoring stations are generally used to assess 

exposure. Because of the logistical burden associated with direct data collection from individual 

subjects, field/panel studies tend to be small in both numbers of subjects and in duration of 

follow-up. While this may limit the statistical power of field studies as compared with timeseries studies, the ability to determine individual-level information on health outcomes and 

potentially confounding factors adds scientific value. 

The most common outcomes measured in acute field studies on the effects of air pollution 

exposure are lung function and various respiratory symptoms. Other respiratory outcomes 

examined on a limited basis include inflammation and generation of hydroxyl radicals in the 

upper airways, as well as numbers of school absences. Also, several studies examined 

cardiovascular outcomes, including heart rate variability (HRV) and risk of myocardial 

infarctions (MI). The first group of studies provided varying degrees of evidence supporting the 

conclusion that elevated O3 levels have negative impacts on lung function and symptoms, 

confirming and adding to the body of knowledge that was presented in the 1996 O3 AQCD. 

Some emphasis has been placed in examining the independent role of O3 in the presence of PM 

and other pollutants. The other new studies contribute information regarding possible O3-related 

cardiopulmonary outcomes that have not previously been as well documented. 

7.2.3 Effects of Acute Ozone Exposure on Lung Function 

As discussed in the 1996 O3 AQCD and in the earlier chapter of this document on 

controlled human exposure studies (Chapter 6), a large body of literature from clinical and field 

studies has clearly and consistently demonstrated reversible decrements in pulmonary function 

following acute O3 exposure. Significant O3-induced spirometric and symptom responses have 

been observed in clinical studies of exercising healthy young adults (see Section 6.2) and in 

some potentially susceptible subpopulations, namely asthmatics and children (see Sections 6.3.2 

and 6.5.1). Field studies of acute O3 exposure that examine pulmonary function fall into two 

distinct groupings, those that conduct spirometry (measuring FEV1, FVC [forced vital capacity], 

and other spirometric indices) and those that measure PEF using peak flow meters. Results from 

the previous O3 AQCD and Chapter 6 of this document support the conclusion that the 

spirometric parameter, FEV1, is a strong and consistent measure of lung function and may be 

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used in the assessment of asthma (Fuhlbrigge et al., 2001). PEF is a closely related but different 

metric of lung function. PEF measurements have been shown to be more variable than FEV1 in 

some studies (Vaughan et al., 1989; Cross and Nelson, 1991) and can have an element of 

uncertain reliability when self-administered by study subjects. However, Lippmann and Spektor 

(1998) state that PEF measurements from small inexpensive flow meters, which are more 

feasible to use in field studies, can produce similar results to PEF measured spirometrically. 

Studies of FEV1 are assessed here first, followed by a discussion of PEF studies. Other 

dividing aspects within these two major types of lung function studies include health status of 

subjects (e.g., healthy, mildly asthmatic, severely asthmatic), age group, time spent outdoors, 

and exertion levels. Several studies brought these factors together to produce informative data. 

Some FEV1 studies involved both increased outdoor O3 exposure and higher exertion levels. 

The results from this group of subjects are relatively comparable to those seen with exercising 

subjects in the clinical studies discussed in Chapter 6. 

7.2.3.1 Spirometry (FEV1) Studies in Outdoor Worker, Exercise, Children, Elderly, and 

Asthmatic Panels 

Studies published over the past decade have provided some new insights on the acute 

effects of O3 on FEV1. The results of all studies that investigated quantitative O3-related effects 

on FEV1 are summarized in the following tables. Tables 7-1a,b,c present changes in FEV1 

associated with O3 exposure in adults, whereas Tables 7-2a,b,c present effects in children. 

Borsboom et al. (1999) examined the circadian variation in spirometric parameters and observed 

that FEV1, FVC, and PEF, in general, increased from early morning until around noon, then 

decreased afterwards. Average variations in FEV1, FVC, and PEF were 2.8%, 4.8%, and 3.1%, 

respectively. To take into account the circadian variation, several studies stratified their analyses 

by time of day, examining the effect of O3 on lung function separately for morning and afternoon 

measurements. Results from these studies are shown in Tables 7-1b and 7-2b. In other 

studies, O3 exposure was related to the cross-day change in spirometric measurements (i.e., the 

difference between same day morning and afternoon measurements). Results from these studies 

are presented in Tables 7-1c and 7-2c. Studies that did not provide quantitative O3 data (Cuijpers 

et al., 1994; Delfino et al., 2004; Frischer et al., 1997) are not included in the tables. The data 

presented in Höppe et al. (1995a) were further analyzed in a subsequent paper (Höppe et al., 

2003); results from the latter paper are included in the tables. In general, the O3 effect estimates 

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Table 7-1a. Field Studies that Investigated the Association Between Acute Ambient O3 

Exposure and Changes in FEV1 in Adults 

Mean O3 (SD) 

Reference Study Location Study Period Level, ppb O3 Index 

Korrick et al. (1998) Mount Washington, NH Summers 1991, 1992 40 (12) 8-h avg a 

Brauer et al. (1996) Fraser Valley, British Jun-Aug 1993 40.3 (15.2) 1-h max 

Columbia, Canada 

Schindler et al. (2001) Eight communities in 

Switzerland 

May-Sep 1991 46.6 

(1.5-127.6) b 

8-h avg 

Höppe et al. (2003) Munich, Germany Apr-Sep 1992-1995 65.9 - 70.4 c 1⁄2-h max 

Romieu et al. (1998) Mexico City Mar-May 1996; 123 (40) 1-h max 

Jun-Aug 1996 

a

 Average of the hourly O3 concentrations during each hike. Hikes averaged 8 hours in duration. b

 Range of 8-h avg concentrations is presented by Schindler et al. (2001). c

 Range of mean 1⁄2-h max O3 concentrations on high O3 days is presented for Höppe et al. (2003).

showed decrements for FEV1 across studies, especially in children. The studies presented in the 

tables are discussed in further detail below, starting with O3 effects on individuals with elevated 

exertion levels and increased exposure due to time spent outdoors, followed by O3 effects on 

persons in other potential risk groups. 

Outdoor worker and exercise panels 

A very important part of the basis for the current 8-h NAAQS for O3 was the results from 

controlled human exposure studies, as discussed in Chapter 6. These field studies with subjects 

at elevated exertion levels are of particular interest due to their similarities to the human 

chamber studies. The majority of human chamber studies have examined the effects of O3 

exposure in subjects performing continuous or intermittent exercise for variable periods of time 

(see Chapter 6 of this document). 

A study by Brauer et al. (1996) reported unusually large O3 effects on lung function among 

outdoor workers. This study presented O3 effects observed during an extended outdoor exposure 

period combined with elevated levels of exertion. The investigators repeatedly measured 

spirometric lung function before and after outdoor summer work shifts over 59 days among a 

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Table 7-1b. Percent Changes in FEV1 (95% CI) Associated with Acute Ambient O3 

Exposures in Adults, Ordered by Size of the Estimate a 

Reference Study Population/Analysis N % Change in FEV1

 1 Brauer et al. (1996) b Berry pickers, next morning 58 !6.36 (!8.02, !4.70)

 2 Brauer et al. (1996) b Berry pickers, afternoon 58 !5.40 (!6.51, !4.28)

 3 Romieu et al. (1998) c Street workers on placebo, 1st phase (lag 0-1) 19 !3.55 (!6.28, !0.82)

 4 Schindler et al. (2001) Adults who never smoked (lag 0) 3,912 !2.96 (!5.11, !0.76)

 5 Romieu et al. (1998) c Street workers on placebo, 1st phase (lag 0) 19 !2.17 (!3.45, !0.89)

 6 Höppe et al. (2003) b Athletes, afternoon (lag 0) 43 !1.26 (!2.63, 0.10)

 7 Romieu et al. (1998) c Street workers on supplement, 1st phase 22 !1.25 (!4.36, 1.86)

(lag 0-1) 

8 Romieu et al. (1998) c Street workers on supplement, 1st phase (lag 0) 22 !0.53 (!2.08, 1.01)

 9 Romieu et al. (1998) c Street workers on placebo, 2nd phase (lag 0) 23 !0.40 (!1.94, 1.14) 

10 Romieu et al. (1998) c Street workers on placebo, 2nd phase (lag 0-1) 23 !0.36 (!2.93, 2.20) 

11 Höppe et al. (2003) Elderly, morning (lag 2) 41 !0.22 (!3.86, 3.42) 

12 Höppe et al. (2003) Athletes, morning (lag 0) 43 !0.01 (!0.10, 0.09) 

13 Romieu et al. (1998) c Street workers on supplement, 2nd phase (lag 0) 19 0.18 (!0.72, 1.08) 

14 Höppe et al. (2003) b Athletes, afternoon (lag 2) 43 0.24 (!0.64, 1.12) 

15 Höppe et al. (2003) b Athletes, afternoon (lag 1) 43 0.48 (!0.97, 1.94) 

16 Höppe et al. (2003) b Athletes, morning (lag 2) 43 0.62 (!0.45, 1.68) 

17 Höppe et al. (2003) b Athletes, morning (lag 1) 43 0.71 (!0.65, 2.07) 

18 Höppe et al. (2003) Elderly, afternoon (lag 0) 41 0.75 (!2.08, 3.58) 

19 Romieu et al. (1998) c Street workers on supplement, 2nd phase 19 0.82 (!0.77, 2.42) 

(lag 0-1) 

20 Höppe et al. (2003) Elderly, afternoon (lag 1) 41 1.16 (!1.26, 3.58) 

21 Höppe et al. (2003) Elderly, morning (lag 0) 41 1.68 (!3.72, 7.07) 

22 Höppe et al. (2003) Elderly, morning (lag 1) 41 1.82 (!2.19, 5.84) 

23 Höppe et al. (2003) Elderly, afternoon (lag 2) 41 2.88 (!0.24, 6.00) 

a

Change in FEV1 is per standard unit ppb O3 (40 ppb for 1⁄2-h max O3 and 1-h max O3, 30 ppb for 8-h max O3,

 and 20 ppb for 24-hr avg O3). b

Brauer et al. (1996) and Höppe et al. (2003) studies also included children. The study population for

 Brauer et-al. ranged in age from 10 to 69 years (mean age 44 years). For Höppe et al. (2003), the athletes

 ranged in age from 13 to 38 years (mean age 18 years). 

c

Romieu et al. (1998) present change in FEV1 (mL). The data from Romieu et al. (1998) were transformed

 to percent change by dividing the estimates by 3,300 mL (average FEV1 for 40 year old Mexican-American

 males by Hankinson et al., 1999). 

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Table 7-1c. Cross-day Percent Changes in FEV1 (95% CI) Associated with Acute 

Ambient O3 Exposures in Adults, Ordered by Size of the Estimate a 

Cross-day % Change 

Reference Study Population/Analysis N in FEV1 

1 Korrick et al. (1998) Hikers with wheeze or asthma (post!pre-hike) 40 !4.47 (!7.65, !1.29) 

2 Korrick et al. (1998) Hikers who hiked 8-12 hours (post!pre-hike) 265 !2.07 (!3.78, !0.36) 

3 Korrick et al. (1998) Hikers age 28-37 years (post!pre-hike) 185 !2.01 (!3.42, !0.60) 

4 Korrick et al. (1998) Hikers who never smoked (post!pre-hike) 405 !1.77 (!3.24, !0.30) 

5 Korrick et al. (1998) Hikers male (post!pre-hike) 375 !1.65 (!3.12, !0.18) 

6 Korrick et al. (1998) Hikers age 38-47 years (post!pre-hike) 142 !1.59 (!3.12, !0.06) 

7 Korrick et al. (1998) All hikers (post!pre-hike) 530 !1.53 (!2.82, !0.24) 

8 Korrick et al. (1998) All hikers, with PM2.5 and acidity in model 530 !1.44 (!3.32, 0.44) 

(post!pre-hike) 

9 Korrick et al. (1998) Hikers age 18-27 years (post!pre-hike) 135 !1.29 (!2.88, 0.30) 

10 Korrick et al. (1998) Hikers female (post!pre-hike) 155 !1.17 (!3.46, 1.12) 

11 Korrick et al. (1998) Hikers age 48-64 years (post!pre-hike) 68 !1.14 (!3.08, 0.80) 

12 Korrick et al. (1998) Hikers without wheeze or asthma 490 !1.08 (!2.49, 0.33) 

(post!pre-hike) 

13 Korrick et al. (1998) Hikers who hiked 2-8 hours (post!pre-hike) 265 !0.99 (!2.70, 0.72) 

14 Korrick et al. (1998) Hikers who formerly smoked (post!pre-hike) 125 !0.72 (!3.07, 1.63) 

15 Brauer et al. (1996) b Berry pickers (post!pre-work shift) 58 0.00 (!1.66, 1.66) 

a

Change in FEV1 is per standard unit ppb O3 (40 ppb for 1⁄2-h max O3 and 1-h max O3, 30 ppb for 8-h max O3,

 and 20 ppb for 24-h avg O3). b

Brauer et al. (1996) study also included children. The study population ranged in age from 10 to 69 years

 (mean age 44 years).

group of 58 berry pickers in Fraser Valley, British Columbia, Canada. The subjects, both male 

and female native Punjabi-speakers, ranged in age from 10 to 69 years old, with a mean age of 

44 years. Outdoor work shifts averaged 11 hours in duration. The mean 1-h max O3 

concentration was 40.3 ppb (SD 15.2). Exertion levels were estimated using portable heart rate 

monitors carried over a period of four or more hours by a representative subset of subjects 

during 16 work shifts. Heart rates over the work shift averaged 36% higher than resting levels. 

Post-shift FEV1 and FVC decreased as a function of O3 concentration, and the effects of O3 

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Table 7-2a. Field Studies that Investigated the Association Between Acute Ambient O3 

Exposure and Changes in FEV1 in Children 

Mean O3 (SD) 

Reference Study Location Study Period Level, ppb O3 Index 

Linn et al. (1996) Rubidoux, Upland, Fall-spring 23 (12) 24-h avg 

and Torrance, CA 1992-1993, 

1993-1994 

Scarlett et al. (1996) Surrey, England Jun-Jul 1994 50.7 (24.48) 8-h max 

Höppe et al. (2003) Munich, Germany Apr-Sep 1992-1995 65.9 - 70.4 a 1⁄2-h max 

Ulmer et al. (1997) Freudenstadt and 

Villingen, Germany 

Mar-Oct 1994 Freudenstadt: 

50.6 (22.5-89.7) b 

1⁄2-h max 

1⁄2-h max 

Villingen: 

32.1 (0.5-70.1) b 

Castillejos et al. (1995) SW Mexico City Aug 1990-Oct 1991 112.3 (0-365) c 1-h max 

Romieu et al. (2002) Mexico City Oct 1998-Apr 2000 102 (47) 1-h max 

Chen et al. (1999) Sanchun, Taihsi, and May 1995-Jan 1996 19.7 - 110.3 c 1-h max 

Linyuan, Taiwan 

a

Range of mean 1⁄2-h max O3 concentrations on high O3 days is presented for Höppe et al. (2003). b

Median and 90th percentile interval are presented for Ulmer et al. (1997).

c

Range of 1-h max O3 concentrations are presented by Castillejos et al. (1995) and Chen et al. (1999).

remained significant after adjusting for PM2.5 in the analysis (see Table 7-1b). Declines in 

lung function also were observed on the morning following high O3 exposure. Ozone was not 

significantly associated with a cross-day change in lung function (i.e., difference between 

afternoon and morning spirometry measurements) (see Table 7-1c). One explanation for the 

lack of an association may be that each spirometric measurement incorporated the impact of the 

previous day’s O3 in addition to the O3 effect on that particular day. The effects seen in this 

study are larger than have been reported previously in studies with briefer exposure durations. 

For example, a change of !3.8 mL (95% CI: !4.6, !3.0) in afternoon FEV1 was shown per 

1 ppb increase in O3 concentrations, compared to the decline of 0.4 mL/ppb and 1.35 mL/ppb 

observed in the earlier adult exercise studies (Spektor et al., 1988b; Selwyn et al., 1985). These 

results are consistent with the interpretation that extended exposures to O3 produce more marked 

effects on lung function. Further, when data were restricted to days with 1-h max O3 

concentrations under 40 ppb, the O3 effects on afternoon FEV1 did not change in magnitude and 

remained significant. However, a possible role of copollutants cannot be completely excluded. 

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Table 7-2b. Percent Changes in FEV1 (95% CI) Associated with Acute Ambient O3 

Exposures in Children, Ordered by Size of the Estimate a 

Reference Study Population/Analysis N % Change in FEV1 

1 Ulmer et al. (1997) b School children in Freudenstadt (lag 1) 57 !4.60 (!7.54, !1.67) 

2 Ulmer et al. (1997) b School boys in Freudenstadt and Villingen 67 !3.23 (!6.47, 0.00) 

(lag 1) 

3 Ulmer et al. (1997) b School children in Freudenstadt and Villingen 135 !2.98 (!5.33, !0.63) 

(lag 1) 

4 Ulmer et al. (1997) b School girls in Freudenstadt and Villingen 68 !2.32 (!5.53, 0.88) 

(lag 1) 

5 Höppe et al. (2003) c Asthmatics, afternoon (lag 2) 43 !2.08 (!6.24, 2.08) 

6 Chen et al. (1999) Children, with NO2 in model (lag 1) 895 !1.97 (!3.51, !0.43) 

7 Chen et al. (1999) Children (lag 1) 895 !1.48 (!2.84, !0.12) 

8 Höppe et al. (2003) Children, morning (lag 0) 44 !1.45 (!4.27, 1.38) 

9 Romieu et al. (2002) b Moderate to severe asthmatic children on 35 !0.99 (!1.80, !0.18) 

placebo (lag 1) 

10 Romieu et al. (2002) b Moderate to severe asthmatic children on 35 !0.97 (!1.87, !0.07) 

placebo, with NO2 and PM10 in model (lag 1) 

11 Chen et al. (1999) Children (lag 2) 895 !0.93 (!2.56, 0.71) 

12 Ulmer et al. (1997) b School children in Villingen (lag 1) 78 !0.79 (!3.93, 2.34) 

13 Chen et al. (1999) Children (lag 7) 895 !0.72 (!1.81, 0.37) 

14 Höppe et al. (2003) c Asthmatics, afternoon (lag 1) 43 !0.56 (!4.61, 3.50) 

15 Linn et al. (1996) b School children, next morning 269 !0.27 (!0.79, 0.24) 

16 Linn et al. (1996) b School children, afternoon 269 !0.19 (!0.73, 0.35) 

17 Romieu et al. (2002) b All asthmatic children on placebo (lag 1) 78 !0.19 (!0.71, 0.33) 

18 Höppe et al. (2003) Children, afternoon (lag 0) 44 !0.14 (!2.71, 2.42) 

19 Höppe et al. (2003) c Asthmatics, afternoon (lag 0) 43 !0.10 (!6.59, 6.39) 

20 Romieu et al. (2002) b Moderate to severe asthmatic on supplement 47 !0.04 (!0.80, 0.72) 

(lag 1) 

21 Romieu et al. (2002) b Moderate to severe asthmatic on supplement, 47 !0.01 (!0.82, 0.80) 

with NO2 and PM10 in model (lag 1) 

22 Scarlett et al. (1996) d School children (lag 1) 154 0.01 (!0.20, 0.22) 

23 Romieu et al. (2002) b All asthmatic children on supplement (lag 1) 80 0.04 (!0.52, 0.60) 

24 Höppe et al. (2003) c Asthmatics, morning (lag 1) 43 0.30 (!3.93, 4.53) 

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Table 7-2b (cont’d). Percent Changes in FEV1 (95% CI) Associated with Acute Ambient 

O3 Exposures in Children, Ordered by Size of the Estimate a 

Reference Study Population/Analysis N % Change in FEV1 

25 Höppe et al. (2003) Children, morning (lag 1) 44 0.83 (!0.53, 2.20) 

26 Höppe et al. (2003) Children, afternoon (lag 1) 44 0.93 (!0.80, 2.66) 

27 Höppe et al. (2003) Children, morning (lag 2) 44 1.17 (!0.36, 2.70) 

28 Höppe et al. (2003) Children, afternoon (lag 2) 44 1.20 (!0.12, 2.52) 

29 Höppe et al. (2003) c Asthmatics, morning (lag 2) 43 1.40 (!3.69, 6.49) 

30 Höppe et al. (2003) c Asthmatics, morning (lag 0) 43 3.41 (!2.50, 9.33) 

a

Change in FEV1 is per standard unit ppb O3 (40 ppb for 1⁄2-h max O3 and 1-h max O3, 30 ppb for 8-h max O3,

 and 20 ppb for 24-h avg O3). b

Linn et al. (1996), Romieu et al. (2002), and Ulmer et al. (1997) present change in FEV1 (mL). The data were

 transformed to percent change by dividing the estimates by 1,900 mL (average FEV1 among 8 to 10 year olds

 by Hankinson et al., 1999). 

c

Höppe et al. (2003) study also included young adults. The study population age for the asthmatics ranged 

 from 12 to 23 years (mean age 15 years). 

d

FEV0.75 results are presented in Scarlett et al. (1996). 

Table 7-2c. Cross-day Percent Changes in FEV1 (95% CI) Associated with Acute 

Ambient O3 Exposures in Children, Ordered by Size of the Estimate a 

Cross-day % Change 

Reference Study Population/Analysis N in FEV1 

1 Linn et al. (1996) b School children (p.m.!a.m.) 269 !0.61 (!1.09, !0.14) 

2 Castillejos et al. (1995) Private primary school (post!pre-exercise) 40 !0.48 (!0.72, !0.24) 

a

Change in FEV1 is per standard unit ppb O3 (40 ppb for 1⁄2-h max O3 and 1-h max O3, 30 ppb for 8-h max O3,

 and 20 ppb for 24-h avg O3). b

Linn et al. (1996) present change in FEV1 (mL). The data were transformed to percent change by dividing the

 estimates by 1,900 mL (average FEV1 among 8 to 10 year olds by Hankinson et al., 1999).

In a Mexico City study of 47 outdoor street workers (Romieu et al., 1998), spirometry was 

performed repeatedly at the end of the work shift over a 2-month period. Subjects were exposed 

to outdoor ambient O3 for a mean of 7.4 h during the workday. The mean 1-h max O3 

concentration during the study period was 123 ppb (SD 40). Among those who had never taken 

an antioxidant supplement (subjects who received a placebo during the first phase of the study), 

7-35
same day O3 concentrations were associated with decreases in FEV1. A mean change of 

!71.6 mL (95% CI: !113.9, !29.3) (approximately a 4% decline) was observed per 40 ppb 

increase in 1-h max O3. The results from this study, in addition to those from the Canadian study 

of berry pickers (Brauer et al., 1996), indicate that outdoor workers are a potentially vulnerable 

population that may need protection from O3 exposures. 

Höppe et al. (1995a) examined forestry workers (n = 41) for changes in pulmonary 

function attributable to O3 exposure in Munich, Germany. In addition, athletes (n = 43) were 

monitored in the afternoon following a 2-h outdoor training period. Pulmonary function tests 

were conducted on days of both “high” (mean 1⁄2-h max O3 of 64 to 74 ppb) and “low” (mean 

1⁄2-h max O3 of 32 to 34 ppb) ambient O3 concentrations. Ventilation rates were estimated from 

the average activity levels. Athletes, who had a fairly high ventilation rate of 80 L/min, 

experienced a significant decrease of 60.8 mL (95% CI: 6.4, 115.2) in FEV1 per 40 ppb increase 

in mean 1⁄2-h max O3. Among the forestry workers, a similar O3-related decline in FEV1 also was 

observed (!56.0 mL [95% CI: !118.4, 6.4]). In a subsequent study, Höppe et al. (2003) 

reanalyzed the results of the athletes after stratifying the spirometric data by time of day 

(morning versus afternoon) and at different lag periods (lags of 0 to 2 days). The reanalysis 

indicated that O3-related decrements were observed only with the afternoon FEV1 at a 0-day lag, 

!1.26% (95% CI: !2.63, 0.10) change in FEV1 per 40 ppb increase in mean 1⁄2-h max O3. The 

mean 1⁄2-h max O3 in the afternoon (1 p.m. to 4 p.m.) was 65.9 ppb (range 51 to 86) on O3 days. 

One FEV1 study clearly demonstrated small but measurable effects of multihour O3 

exposures on adults exercising outdoors. In Korrick et al. (1998), adult hikers (n = 530) of 

Mount Washington, NH performed spirometry before and after hiking for a mean of 8 hours 

(range 2 to 12). The mean hourly O3 concentration during each hike ranged from 21 to 74 ppb. 

After the hike, all subjects combined experienced a small mean decline of 1.5% (95% CI: 0.2, 

2.8) in FEV1 and 1.3% (95% CI: 0.5, 2.1) in FVC per 30 ppb increase in the mean of the hourly 

O3 concentration during the hike. In addition, Korrick et al. (1998) compared hikers who hiked 8 

to 12 hours to those who hiked 2 to 8 hours. Among those who hiked longer, the percent change 

in FEV1 was more than 2-fold greater per ppb exposure compared to those who hiked only for 2 

to 8 hours. Each hour hiked, which may reflect dose, was associated with a decline of 0.3% 

(p = 0.05) in FEV1, after adjusting for O3. 

7-36
In a Mexico City study, the O3 effect attributable to exercise was determined using a group 

of school children (n = 40) who were chronically exposed to moderate to high levels of O3 

(Castillejos et al., 1995). Children were tested up to 8 times between August 1990 and October 

1991. Spirometry was performed by the children before and after a 1-h intermittent exercise 

session outdoors. Outdoor O3 levels ranged up to 365 ppb, with a mean of 112.3 ppb. Linear 

trend analyses indicated a relationship between quintiles of O3 and percent change in lung 

function. However, stratified analyses indicated that significant changes were observed only 

with higher quintiles of O3 exposure (72 to 125 ppb and 183 to 365 ppb). Therefore, children 

exercising at higher O3 levels experienced declines in pulmonary function despite the repeated 

daily exposure to moderate and high levels of O3 in Mexico City. 

Collectively, the above studies confirm and extend clinical observations that prolonged 

exposure periods, combined with elevated levels of exertion or exercise, may magnify the effect 

of O3 on lung function. The most representative data come from the Korrick et al. (1998) hiker 

study. This U.S. study provided outcome measures stratified by several factors (e.g., gender, 

age, smoking status, presence of asthma) within a population capable of more than normal 

exertion. 

Panel studies of children, elderly, and asthmatics 

Höppe et al. (1995a,b) examined several potentially susceptible populations for changes in 

pulmonary function attributable to O3 exposure in Munich, Germany. The forestry workers and 

athletes were discussed in the previous section. Senior citizens (n = 41) and juvenile asthmatics 

(n = 43) were also monitored on “low” O3 (mean 1⁄2-h max O3 of 32 to 34 ppb) and “high” O3 

(mean 1⁄2-h max O3 of 64 to 74 ppb) days. Subjects were requested to stay outdoors for at least 

2 hours just before the afternoon pulmonary function test. Clerks (n = 40) were considered the 

nonrisk control group. Although clerks spent the majority of their time indoors, their outdoor 

exposures on high O3 days were similar to that of the four other risk groups. The results showed 

no significant O3 effects on the senior citizens. Clinical studies also have consistently shown 

that seniors are less responsive to O3 (Bedi et al., 1989; Drechsler-Parks, 1995). Asthmatics and 

clerks experienced slight reductions in FEV1 on high O3 days. Among all risk groups, juvenile 

asthmatics experienced the largest O3-related decline in FEV1, !84.0 mL (95% CI: !196.4, 

28.4) per 40 ppb increase in mean 1⁄2-h max O3. To further examine their hypotheses on 

7-37
characteristics of O3 risk groups, Höppe et al. (2003) conducted a different analysis on a more 

expanded data base than utilized in the earlier study. Children were examined as an additional 

risk group. Mean 1⁄2-h max O3 ranged from 65.9 ppb to 70.4 ppb on O3 days for each risk group. 

Höppe et al. (2003) presented both group mean values and analyses on an individual basis. For 

the group mean values, consistent O3 effects were not detectable. On an individual basis, a 

potential pattern of O3 sensitivity was observed (see Table AX7-1 in Annex Section AX7.1 for 

details). About 20% of the children and asthmatics were regarded as O3 responders (i.e., 

individuals that had greater than 10% change in FEV1) compared to only 5% of the elderly and 

athletes. These results indicated that while the majority of the population did not react to O3 

exposure, a small group of susceptible individuals experienced health effects from O3. The 

sample size limits quantitative extrapolation to larger populations, but may allow cautious first 

estimates. 

Several other panel studies performed spirometry in children, another potentially 

susceptible group (Avol et al., 1998; Chen et al., 1999; Cuijpers et al., 1994; Frischer et al., 

1997; Linn et al., 1996; Romieu et al., 2002; Scarlett et al., 1996; Ulmer et al., 1997). 

All studies, with the exception of Avol et al. (1998) and Scarlett et al. (1996), observed a 

decrease in FEV1 associated with O3 exposure. In a cohort of 154 children in Surrey, England, 

Scarlett et al. (1996) observed no association between ambient O3 concentrations and FEV0.75 

(0.2 mL [95% CI: !3.6, 3.9] increase per 30 ppb increase in 8-h max O3 at a 1-day lag), but 

noted a small effect of PM10 on lung function. The mean 8-h max O3 concentration was 50.7 ppb 

(range 6.8 to 128 ppb). The study by Avol et al. (1998) examined three groups of children, 

asthmatic (n = 53), wheezy (n = 54), and healthy (n = 103). Ozone levels were reported as being 

very low (values not provided). The authors advised that noncompliance by the subjects might 

have been a problem, and further noted limitations in the analysis methods and other aspects of 

the study design. 

One large study measured spirometric lung function in 895 school children in three towns 

in Taiwan (Chen et al., 1999). The 1-h max O3 concentrations ranged from 19.7 to 110.3 ppb. 

Lung function was measured only once for each subject. The authors reported significant 

associations between diminished FEV1 and FVC with a 1-day lag of O3 concentrations. Effect 

sizes were typical of those observed in past studies, i.e., 0.5 to 1.0 mL decline in FEV1 per ppb 

7-38
increase in O3 concentration. Ozone was the only air pollutant associated with changes in lung 

function in multipollutant models including SO2, CO, PM10, and NO2. 

Linn et al. (1996) repeatedly measured spirometric lung function among 269 school 

children in three southern California communities (Rubidoux, Upland, and Torrance). Lung 

function was measured over 5 consecutive days, once in each of 3 seasons over 2 school years. 

Between-week variability was controlled in the analysis by seasonal terms in the model. 

Statistical power was limited by the relatively narrow range of exposures that were experienced 

within each week. In addition, the study was restricted to the school year, eliminating most of 

the “high” O3 season from consideration. During the study period, 24-h avg O3 levels at the 

central monitoring site ranged up to 53 ppb (mean 23 ppb), whereas personal measurements 

ranged up to 16 ppb (mean 5 ppb). A mean change of !11.6 mL (95% CI: !20.6, !2.6) 

(approximately a 1% decline) in FEV1 was observed from morning to afternoon per 20 ppb 

increase in 24-h avg O3. Other associations (involving individual morning or afternoon FVC 

and FEV1 measurements) went in the plausible direction, but the O3 effect estimates were 

much smaller. 

Ulmer et al. (1997) examined 135 children aged 8 to 11 years in two towns in Germany 

from March to October 1994 for O3 effects on pulmonary function at four time periods. The 

cross-sectional results at each of the four time points showed limited FVC and no FEV1 

associations. However, the longitudinal analysis, which combined data from all four periods 

yielded a mean change of !87.5 mL (95% CI: !143.2, !31.7) (approximately a 5% decline) 

in FEV1 per 40 ppb increase in 1⁄2-h max O3 for the town with the higher O3 levels (median 1⁄2-h 

max of 50.6 ppb versus 32.1 ppb). In the cross-sectional analysis, only between-person 

variability was analyzed. The longitudinal analysis, in which the subjects provided multiple 

days of measurements, provided information on both between- and within-subject responses. 

There are a limited number of new epidemiologic studies examining the effects of O3 

on FEV1; however, results from these studies indicate that acute exposure to O3 is associated 

with declines in FEV1 in children. These results further support the negative effects of O3 on 

lung function observed in the meta-analysis on children attending summer camp (Kinney et al., 

1996a) and in the clinical literature. 

7-39
7.2.3.2 Peak Flow Meter (PEF) Studies in Asthmatics and Healthy Individuals 

Many studies of the acute effect of O3 on PEF examined self-administered PEF levels 

daily, both in the morning and afternoon. PEF follows a circadian rhythm, with the highest 

values found during the afternoon and lowest values during the night and early morning 

(Borsboom et al., 1999). Due to the diurnal variation in PEF, most studies analyzed their data 

after stratifying by time of day. The peak flow studies examined both asthmatic panels and 

healthy individuals. The asthma panels are discussed first. 

Asthma panels 

The effects of acute O3 exposure on PEF in asthmatics were examined in several panel 

studies. Figures 7-1 and 7-2 present percent changes in morning and afternoon PEF outcomes 

from seven panel studies of children, mostly asthmatic, ranging in age from 5 to 13 years. The 

effect estimates from all single-day and multiday lag models are presented. Only single-city 

results with analyses stratified by morning and afternoon are included in the figure. Studies that 

examined cross-day changes and daily variability in PEF (e.g., Just et al., 2002; Thurston et al., 

1997) are not included in the figure since such outcomes are not directly comparable. 

Collectively, nearly all of the studies indicated decrements of peak flow but most of the 

individual estimates were not statistically significant. The results from the individual studies are 

further discussed below. 

In Mexico City, two studies of asthmatic school children were carried out simultaneously 

in the northern (Romieu et al., 1996) and southwestern sections of the city (Romieu et al., 1997). 

In the northern study, 71 mildly asthmatic school children aged 5 to 13 years old, were followed 

over time for daily morning (before breakfast) and afternoon (bedtime) PEF. The mean 1-h 

max O3 was 190 ppb (SD 80). In single-pollutant models, O3 concentrations at 0-, 1-, and 2-day 

lags were associated with diminished morning and afternoon PEF, but only the 0-day lag 

morning effect was significant. The O3 effect became nonsignificant when PM2.5 was added to 

the model. In the southwestern study, 65 mildly asthmatic children aged 5 to 13 years old were 

followed during the summer and winter for daily morning and afternoon PEF. The mean 1-h 

max O3 was 196 ppb (SD 78). Ozone concentrations at a 0- and 1-day lag were associated with 

afternoon PEF, with larger effects at a 1-day lag. Associations involving O3 were stronger than 

7-40
Figure 7-1. Percent change (95% CI) in morning PEF in children per standardized 

increment (see Section 7.1.3.2). For single-day lag models, previous day O3 

effects are shown. For multiday lag models, the cumulative effects of a 1- to 

5-day lag are shown for Mortimer et al. (2002) and Neas et al. (1999), and the 

effect of a 1- to 10-day lag is shown for Gold et al. (1999). 

those involving PM10. Several additional studies, both in the United States and in other 

countries, reported significant associations between O3 exposure and decrements in PEF among 

asthmatics (Gielen et al., 1997; Jalaludin et al., 2000; Just et al., 2002; Ross et al., 2002; 

Thurston et al., 1997). 

7-41
Figure 7-2. Percent change (95% CI) in afternoon PEF in children per standardized 

increment (see Section 7.1.3.2). For single-day lag models, current day O3 

effects are shown. For multiday lag models, the cumulative effect of a 1- to 

5-day lag is shown for Neas et al. (1999) and a 1- to 9-day lag is shown for 

Gold et al. (1999). 

Other epidemiologic studies did not find a significant O3 effect on the lung function of 

asthmatics. Delfino et al. (1997a) examined morning and evening PEF among 22 asthmatics 

ranging in age from 9 to 46 years, living in Alpine, CA. Daily ambient 12-h avg O3 (8 a.m. to 

8 p.m.) concentrations ranged from 34 to 103 ppb, with a mean value of 64 ppb. Unique to this 

study, personal O3 exposures were measured using 12-h passive O3 samplers that were worn by 

7-42
the subjects. The personal 12-h avg O3 (8 a.m.-8 p.m.) concentrations, which had a mean value 

of 18 ppb, were much lower than the fixed-site ambient levels. Quantitative O3 results were not 

reported, but the researchers stated that no O3 effects were observed on morning and evening 

PEF. In Hiltermann et al. (1998), 60 nonsmoking adults aged 18 to 55 years in Bilthoven, the 

Netherlands, were followed between July and October 1995 with morning and afternoon PEF 

measurements. Although negative associations were observed between O3 and cross-day 

changes in PEF, the results were not significant. The mean ambient 8-h max O3 was 41.3 ppb 

(range 3 to 49). 

Mortimer et al. (2002) examined 846 asthmatic children from the National Cooperative 

Inner-City Asthma Study (NCICAS) for O3-related changes in PEF. Children from eight U. S. 

urban areas (St. Louis, MO; Chicago, IL; Detroit, MI; Cleveland, OH; Washington, DC; 

Baltimore, MD; East Harlem, NY; and Bronx NY) were monitored from June through August 

1993. Median 8-h avg O3 (10 a.m.-6 p.m.) concentrations ranged from 34 ppb in Chicago to 

58 ppb in Washington, DC. The mean 8-h avg O3 level across the eight cities was 48 ppb. This 

study provides representative data for the United States, in so much as children from multiple 

cities throughout the East and Midwest were examined. Asthmatic children from urban areas are 

an important subgroup of potentially at-risk populations. Study children either had physiciandiagnosed asthma and symptoms in the past 12 months or respiratory symptoms consistent with 

asthma that lasted more than 6 weeks during the previous year. 

Mortimer et al. (2002) examined O3-related changes in PEF for single-day lags from 1 to 

6 days and a multiday lag period of 5 days. Of all the pollutants examined, including O3, PM10, 

NO2, and SO2, none were associated with evening PEF. Only O3 was found to be associated 

with morning PEF. The effect estimates of the association between O3 and morning PEF for the 

single-day and multiday lags are depicted as error density curves in Figure 7-3 (for description of 

error density curves, see Annex Section AX7.2). Small morning effects were observed at 1- and 

2-day lags. The effect of O3 on morning outcomes increased over several days. A strong 

association between O3 and PEF also was found with a multiday lag period (cumulative lag of 

1 to 5 days). Unrestricted lag models suggested that the O3 exposure from 3 to 5 days prior had a 

greater impact on morning % PEF than more immediate exposures. Mortimer et al. discussed 

biological mechanisms for delayed effects on pulmonary function, which included increased 

bronchial reactivity secondary to airway inflammation associated with irritant exposure. Animal 

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Figure 7-3. Percent changes in PEF per 30 ppb increase in 8-h avg O3 in urban children. 

Single-day lags (1-, 2-, 3-, 4-, 5-, and 6-day) are shown in the upper panel. 

The cumulative multiday lag (1- to 5-day) is shown in the lower panel. 

Source: Derived from Mortimer et al. (2002). 

toxicology and human chamber studies (see Chapters 5 and 6) provide further evidence that 

exposure to O3 may augment cellular infiltration and cellular activation, enhance release of 

cytotoxic inflammatory mediators, and alter membrane permeability. 

Figure 7-4 illustrates the probability density curves of the results from the individual-cities 

analysis and that from the pooled analysis of all eight cities. The error density curve for the 

all-cities analysis is a graphical presentation of the all-cities regression analysis presented by 

Mortimer et al. (2002), a change in morning PEF of !1.18% (95% CI: !2.10, !0.26) per 30 ppb 

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Figure 7-4. Percent change in PEF per 30 ppb increase in 8-h avg O3 with a cumulative lag 

of 1 to 5 days. The error density curve is shown for the pooled analysis of the 

eight NCICAS cities and the summary density curve is presented for the 

analyses from the individual cities. Note that 99% and 78% of the areas 

under the curves are less than zero for the pooled-cities analysis and 

individual-cities analysis, respectively. 

Source: Derived from Mortimer et al. (2002). 

increase in 8-h avg O3 with a cumulative lag of 1 to 5 days. The summary density curve for the 

individual-cities analysis was calculated by summing together eight normal distribution 

functions, one for each of the study cities, then taking the derivative of the summed function (see 

Annex Section AX7.2 for further explanation of summary density curves). The area under the 

density curve and to the left of a value on the x-axis is an estimate of the probability that the 

effect estimate will be less than or equal to that value. For example, the area under the density 

curve to the left of 0% change in PEF is 99% in the all-cities analysis. A wider distribution was 

observed in the individual-cities analysis, with only 78% of the area less than zero. The 

summary density curve of the individual-cities analysis has a larger standard error than the error 

density curve of the all-cities analysis, because the summary density does not average the city 

effects. The regression analysis by Mortimer et al. (2002) suggested a lack of heterogeneity by 

city, as indicated by the nonsignificant interaction term between O3 effect and city. As shown in 

Figure 7-4, the summary density curve of the individual-cities analysis has a peak at about the 

same value as the curve of the all-cities analysis, suggesting a common O3 effect for all eight 

7-45
cities and small variation among them. The unimodal shape of the density curve of the 

individual-cities analysis also indicates the absence of outlying cities. 

Mortimer et al. (2002) further noted that small declines in morning PEF may be of 

uncertain clinical significance; thus they calculated the incidence of $10% declines in PEF. 

A 5 to 15% change in FEV1 has been expressed as having clinical importance to asthma 

morbidity (American Thoracic Society, 1991; Lebowitz et al., 1987; Lippmann, 1988). 

Although greater variability is expected in PEF measurements, a $10% change in PEF also may 

have clinical significance. In Mortimer et al. (2002), O3 was associated with an increased 

incidence of $10% declines in morning PEF (odds ratio of 1.30 [95% CI: 1.04, 1.61] per 30 ppb 

increase in 8-h avg O3 for a 5-day cumulative lag). This finding suggests that exposure to O3 

might be related to clinically important changes in PEF in asthmatic children. This study also 

observed that excluding days when 8-h avg O3 levels were greater than 80 ppb provided effect 

estimates that were similar to those when all days were included in the analysis, indicating that 

the negative effect of O3 on morning PEF persisted at levels below 80 ppb. There is some 

concern, however, regarding the lack of an association between O3 and afternoon PEF. 

Results from the multicities study by Mortimer et al. (2002), as well as those from several 

regional studies, provide evidence of a significant relationship between O3 concentrations and 

PEF among asthmatics. Collectively, these studies indicate that O3 may be associated with 

declines in lung function in this potentially susceptible population. 

Panels of healthy subjects 

The effect of O3 on PEF in healthy subjects also was investigated in several studies. 

A study of 162 children (9 years of age) in England examined the relationship between O3 and 

PEF in the winter and summer seasons (Ward et al., 2002). The median 24-h avg O3 

concentrations were 13.0 ppb in the winter and 22.0 ppb in the summer. The O3 effect estimates 

were generally positive in the winter and negative in the summer. Single-day lags of 0- to 

3-days were examined; however, the strongest association was found with a multiday lag period. 

During the summer, a decline of 11.10 L/min (95% CI: 0.18, 21.98) was observed in morning 

PEF per 20 ppb increase in 24-h avg O3 with a 7-day cumulative lag. Smaller O3 effects were 

observed on afternoon PEF. 

7-46
During the summer of 1990, Neas et al. (1995) examined 83 children in Uniontown, PA 

and reported twice daily PEF measurements. Researchers found that evening PEF was 

associated with O3 levels weighted by hours spent outdoors. Using a similar repeated measures 

design, Neas et al. (1999) saw evidence for effects due to ambient O3 exposure among 

156 children attending two summer day camps in the Philadelphia, PA area. The mean daytime 

12-h avg O3 (9 a.m. to 9 p.m.) levels were 57.5 ppb at the southwestern camp and 55.9 ppb at the 

northeastern camp. Negative associations were found between the preceding 12-h avg O3 and 

afternoon PEF (recorded before leaving camp), as well as morning PEF (recorded upon arrival at 

camp). However, the only significant relationship was between O3 and both morning and 

afternoon PEF considered jointly when a multiday lag period was used. Naeher et al. (1999), in 

a sample of 473 nonsmoking women (age 19 to 43 years) living in Vinton, VA, also showed the 

strongest association between O3 and evening PEF with a 5-day cumulative lag exposure. 

Another study in southwestern Mexico City analyzed morning and afternoon PEF data 

collected from 40 school children aged 8 to 11 years (Gold et al., 1999). Subjects provided 

measurements upon arriving and before departing from school each day. The mean 24-h avg O3 

was 52.0 ppb (IQR 25). A negative effect of O3 on PEF was observed, !1.60 mL/s (95% CI: 

!3.56, 0.36) and !1.80 mL/s (95% CI: !3.76, 0.16) per 20 ppb increase in 24-h avg O3 on the 

same day afternoon and next day morning PEF, respectively. A greater effect was observed for 

PEF regressed on O3 concentrations with a cumulative 10-day lag period (!3.50 mL/s [95% CI: 

!5.52, !1.49] on same day afternoon). These results suggest a longer, cumulative effect of O3 

on PEF. Alternatively, the associations observed at the 10-day lag period may reflect 

confounding by other time-varying factors or be a chance finding from an exploratory analysis. 

In a recent study of 43 mail carriers in Taichung City, Taiwan, PEF was monitored twice 

daily during a six-week period (Chan and Wu, 2005). The mean 8-h avg O3 (9 a.m.-5 p.m.) 

concentration during their work shift was 35.6 ppb (SD 12.1). Associations were observed 

between evening PEF and O3 concentrations at lags of 0, 1 and 2 days. The greatest effect was 

observed at a lag of 1 day, a 2.07% decline in PEF per 30 ppb increase in 8-h avg O3 

(quantitative results for 95% CI not provided). Similar O3 effects on morning PEF were 

observed. The effect of O3 on PEF was robust to adjustment for copollutants; no association 

with PEF was observed for PM10 and NO2 in multipollutant models. 

7-47
7.2.4 Respiratory Symptoms 

Studies published over the past decade represent an improved new body of data on the 

symptom effects of O3. Respiratory symptoms in acute air pollution field studies are usually 

measured using questionnaire forms (or “daily diaries”) that are filled out by study subjects, 

usually without the direct supervision of research staff. Questions address the daily experience 

of coughing, wheezing, shortness of breath (or difficulty breathing), production of phlegm, and 

others. While convenient and potentially useful in identifying acute episodes of morbidity, 

measurements of daily symptoms are prone to a variety of errors. These include, for example, 

misunderstanding of the meaning of symptoms, variability in individual interpretation of 

symptoms, inability to remember symptoms if not recorded soon after their occurrence, reporting 

bias if days of high air pollution levels are identifiable by subjects, and the possibility of falsified 

data. In spite of these potential problems, the ease of data collection has made daily symptom 

assessment a common feature of field studies. Many of the studies reviewed above for lung 

function results also included measurements of daily symptoms. Pearce et al. (1998) reported 

that one advantage associated with the study of asthma panels is that the population is usually 

already familiar with symptom terms, e.g., wheezing and cough. Delfino et al. (1998a) further 

states that the use of repeated daily symptom diaries has additional advantages of reducing recall 

bias, given the proximity of events, and allowing for health effects to be modeled with each 

subject serving as their own control over time. Also, study design can blind the participants 

from the air pollution aspect of the study. Careful efforts by study staff can help ensure that the 

symptom diaries provide information that is less affected by the potential problems noted above. 

Similar to studies of lung function, respiratory symptom studies can be divided into two 

groups, asthma panels or healthy subjects. Asthma panel studies are discussed first. 

Asthma panels 

Most studies examining respiratory symptoms related to O3 exposure focused on asthmatic 

children. Among the health outcomes, of particular interest were those associated with asthma, 

including cough, wheeze, shortness of breath, and increased medication use. Figures 7-5 and 7-6 

present the odds ratios for O3-related cough and medication use among asthmatic children from 

six studies (Gielen et al., 1997; Jalaludin et al., 2004; Just et al., 2002; Ostro et al., 2001; Romieu 

et al., 1996, 1997). For consistency, only single-city or single-region studies that presented odds 

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Figure 7-5. Odds ratios for the incidence of cough among asthmatic children per 

standardized increment (see Section 7.1.3.2). For single-day lag models, 

current day O3 effects are shown with the exception of Ostro et al. (2001), 

which only presented results from a 3-day lag. For multiday lag models, the 

cumulative effects of a 0- to 4-day lag are shown. 

*

Note that Gielen et al. (1997) and Jalaludin et al. (2004) presented results for prevalence of cough. 

7-49 
Figure 7-6. Odds ratios for extra medication use among asthmatic children per 

standardized increment (see Section 7.1.3.2). For single-day lag models, 

current day O3 effects are shown. For multiday lag models, the cumulative 

effects of a 0- to 4-day lag are shown. 

7-50
ratios are included in the figure. Studies that presented change in severity of symptoms or some 

other informative health outcome are not included in the figure, since such symptom outcomes 

differ from indicating simple presence or absence of symptoms. The study by Gent et al. (2003) 

also is not included in this figure, as odds ratios for cough and medication use were analyzed for 

quintiles of O3 concentrations using the lowest quintile as the reference. These studies are 

discussed separately. 

The various effect estimates for the association between O3 concentrations and cough are 

depicted in Figure 7-5. Despite the variability in the individual effect estimates, there is some 

consistency in the O3 effects. In general, the majority of the odds ratios appear to be greater 

than one among the single-day lag models, suggesting an association between acute exposure 

to O3 and increased cough among asthmatic children. Figure 7-6 presents the odds ratios 

for O3-associated bronchodilator use. The results for medication use are less consistent than 

those for cough; one study by Just et al. (2002) observed strong positive associations, but had 

wide confidence intervals. 

Among the studies reporting results for daily symptoms and asthma medication use, 

several observed associations with O3 concentrations that appeared fairly robust (Delfino et al., 

2003; Desqueyroux et al., 2002a,b; Gent et al., 2003; Hiltermann et al., 1998; Just et al., 2002; 

Mortimer et al., 2000, 2002; Newhouse et al., 2004; Romieu et al., 1996, 1997; Ross et al., 2002; 

Thurston et al., 1997). Mortimer et al. (2002) reported morning symptoms in 846 asthmatic 

children from 8 U.S. urban areas to be most strongly associated with a cumulative 1- to 4-day lag 

of O3 concentrations in the NCICAS. The NCICAS used standard protocols that included 

instructing caretakers of the subjects to record symptoms in the daily diary by observing or 

asking the child (Mitchell et al., 1997). Symptoms reported included cough, chest tightness, and 

wheeze. In the analysis pooling individual subject data from all eight cities, the odds ratio for 

the incidence of symptoms was 1.35 (95% CI: 1.04, 1.69) per 30 ppb increase in 8-h avg O3 

(10 a.m.-6 p.m.). The mean 8-h avg O3 was 48 ppb across the 8 cities. Excluding days when 8-h 

avg O3 was greater than 80 ppb (less than 5% of days), the odds ratio was 1.37 (95% CI: 1.02, 

1.82) for incidence of morning symptoms. Figure 7-7 presents the probability density curves of 

the odds ratios for the incidence of symptoms with a 1- to 4-day cumulative lag from the 

individual-cities analysis and the all-cities analysis. This figure confirms the regression results 

indicating that there is a significant increase in odds for incidence of symptoms, as the area 

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Figure 7-7. Odds ratio for the incidence of symptoms per 30 ppb increase in 8-h avg O3 

with a cumulative lag of 1 to 4 days. The error density curve is shown for the 

pooled analysis of the eight NCICAS cities and the summary density curve is 

presented for the analyses from the individual cities. Note that 99% and 76% 

of the areas under the curves are less than zero for the pooled-cities analysis 

and individual-cities analysis, respectively. 

Source: Derived from Mortimer et al. (2002). 

under the density curve with an odds ratio greater than one is 99%. Mortimer et al. (2002) 

did not observe significant interactions among the eight cities, indicating that there was no 

heterogeneity among the city-specific estimates. The unimodal distribution of the city-stratified 

summary density curve also suggests a lack of significant heterogeneity in O3 effects among the 

cities. It should be noted that other pollutants, including PM10 (monitored in 3 cities), NO2 

(in 7 cities), and SO2 (in all 8 cities), also were associated with increased incidence of morning 

symptoms. In multipollutant models, the O3 effect was shown to be slightly diminished. The 

odds ratios for the incidence of symptoms per 30 ppb increase in 8-h avg O3 were 1.23 (95% CI: 

0.94, 1.61) with SO2 and 1.14 (95% CI: 0.85, 1.59) with NO2. In the three urban areas 

with PM10 data, the odds ratios were 1.21 (95% CI: 0.61, 2.40) in the O3-only model and 

1.08 (95% CI: 0.41, 2.40) when PM10 also was included in the model. 

Another one of the larger studies was that of Gent and colleagues (2003), where 

271 asthmatic children under age 12 and living in southern New England were followed over 

6 months (April through September) for daily symptoms in relation to O3 and PM2.5. Mean 1-h 

max O3 and 8-h max O3 concentrations were 58.6 ppb (SD 19.0) and 51.3 ppb (SD 15.5), 

7-52
respectively. The data were analyzed for two separate groups of subjects, 130 who used 

maintenance asthma medications during the follow-up period and 141 who did not. The need for 

regular medication was considered to be a proxy for more severe asthma. Not taking any 

medication on a regular basis and not needing to use a bronchodilator would suggest the 

presence of very mild asthma. Effects of 1-day lag O3 were observed on a variety of respiratory 

symptoms only in the medication user group. Both daily 1-h max and 8-h max O3 concentrations 

were similarly related to symptoms such as chest tightness and shortness of breath. Effects 

of O3, but not PM2.5, remained significant and even increased in magnitude in two-pollutant 

models. Some of the associations were noted at 1-h max O3 levels below 60 ppb. In contrast, no 

effects were observed among asthmatics not using maintenance medication. In terms of persondays of follow-up, this is one of the larger studies currently available that address symptom 

outcomes in relation to O3, and provides supportive evidence for effects of O3 independent 

of PM2.5. Study limitations include limited control for meteorological factors and the post-hoc 

nature of the population stratification by medication use. 

Some international studies have reported significant associations of respiratory symptoms 

with O3. The incidence of asthma attacks was associated with O3 concentrations in a group of 

60 severe asthmatics (mean age 55 years) followed over a 13-month period in Paris, France 

(Desqueyroux et al., 2002a). In a similar study, Desqueyroux et al. (2002b) observed O3­

associated exacerbation of symptoms in 39 adult patients (mean age 67 years) with chronic 

obstructive pulmonary disease (COPD). Interestingly, in contrast to results from controlled 

human exposure studies (see Section 6.3.1, Subjects with COPD), the O3 effect appeared larger 

in this study among subjects who smoked and among those with more severe COPD. However, 

the low O3 concentrations experienced during this study (summer mean 8-h max O3 of 21 ppb 

[SD 9]) raise plausibility questions. In a study of 60 nonsmoking asthmatic adults (aged 18 to 

55 years) in Bilthoven, the Netherlands, Hiltermann and colleagues (1998) reported associations 

between O3 and daily symptoms of shortness of breath and pain upon deep inspiration. The 

mean 8-h max O3 was 41 ppb (range 3 to 49). The O3 associations were stronger than those 

for PM10, NO2, SO2, and black smoke (BS). No differences in response were evident between 

subgroups of subjects defined on the basis of steroid use or airway hyperresponsiveness. Daily 

use of bronchodilators or steroid inhalers was not found to be associated with O3 in this study. 

7-53
Other studies showed only limited or a lack of evidence for symptom increases being 

associated with O3 exposure (Avol et al., 1998; Chen et al., 1998; Delfino et al., 1996, 1997a, 

1998a; Gielen et al., 1997; Jalaludin et al., 2004; Ostro et al., 2001; Taggart et al., 1996). Avol 

et al. (1998) studied symptoms in asthmatic, wheezy, and healthy children aged 10 to 12 years in 

southern California. Some symptom associations were noted but they were inconsistent. For 

example, children with wheeze were at increased risk for difficulty of breathing and for 

wheezing at low O3 concentrations (1-h max O3 of 30 ppb), but not at higher O3 concentrations 

(1-h max O3 of 120 ppb). The authors noted that O3 concentrations were relatively low and that 

children studied did not spend substantial time engaged in physical activities outdoors . Ostro 

et al. (2001) reported no associations between daily symptoms and ambient O3 concentrations in 

a cohort of 138 African-American children with asthma followed over 3 months (August to 

October) in Central Los Angeles and Pasadena, CA. However, the use of extra asthma 

medication was associated with 1-h max O3 concentrations at a 1-day lag. Mean 1-h max O3 

concentrations were 59.5 ppb in Los Angeles and 95.8 ppb in Pasadena. Delfino and colleagues 

(1996) followed 12 asthmatic teens living in San Diego, CA for respiratory symptoms over a 

2-month period and saw no relationship with central site ambient O3 (mean 12-h avg O3 of 

43 ppb [SD 17]). Personal O3 exposures (mean 12-h avg O3 of 11.6 ppb [SD 11.2]) were 

associated with the composite symptom score and $2-agonist inhaler use, but the relationship 

with symptom score disappeared when weekday/weekend differences were controlled in the 

statistical analysis. Study power was likely compromised by the small sample size. This 

observation of stronger associations with O3 levels from personal monitors implies that gains in 

power may be achieved if exposure misclassification is reduced through the use of personal 

exposure measurements rather than central site ambient O3 concentrations. A similar study of 

22 asthmatics in Alpine, CA observed no effects of O3 on symptoms when personal O3 exposure 

(mean 12-h avg O3 of 18 ppb) was used as the exposure metric (Delfino et al., 1997a). However, 

a later study in the same location involving 24 subjects (Delfino et al., 1998a) did find an 

association between respiratory symptoms and ambient O3 concentrations (mean 1-h max O3 of 

90 ppb), with stronger O3 effects experienced by asthmatics not on anti-inflammatory 

medication. In this study, a binary symptom score was used, whereas the earlier study used a 

linear symptom score of 0 through 6. 

7-54
The multicities study by Mortimer et al. (2002), which provides an asthmatic population 

most representative of the United States, and several single-city studies indicate a robust 

association of O3 concentrations with respiratory symptoms and increased medication use in 

asthmatics. However, there are a number of well-conducted, albeit relatively smaller, studies 

that have not found such effects. 

Panels of healthy subjects 

Fewer studies examined the effect of O3 on respiratory symptoms in healthy individuals. 

Neas et al. (1995) reported that, in school children, evening cough was associated with O3 levels 

weighted by hours spent outdoors. The mean daytime 12-h avg O3 (8 a.m.-8p.m.) concentration 

was 50.0 ppb. The study by Linn and colleagues (1996) of 269 school children in southern 

California found no associations between respiratory symptoms and O3, but subjects were 

exposed to fairly low O3 concentrations (mean 24-h avg O3 of 5 ppb [SD 3]) as determined by 

personal monitors. Gold et al. (1999) examined symptoms in 40 healthy children in southwest 

Mexico City. The mean ambient 24-h avg O3 concentration was 52.0 ppb (IQR 25). Pollutant 

exposures were associated with increased production of phlegm in the morning, although the 

effects of the air pollutants (PM2.5, PM10, and O3) could not be separated in multipollutant 

models. Hoek and Brunekreef (1995) did not find a consistent association between ambient O3 

levels and the prevalence or incidence of respiratory symptoms in children living in two rural 

towns in the Netherlands. Mean 1-h max O3 concentrations were 57 ppb (range 22 to 107) 

in Deurne and 59 ppb (range 14 to 114) in Enkhuizen. Collectively, these studies indicate that 

there is no consistent evidence of an association between O3 and respiratory symptoms among 

healthy children. 

7.2.5 Acute Airway Inflammation 

Acute airway inflammation has been shown to occur among adults exposed to 80 ppb O3 

over 6.6 hours with exercise in controlled chamber studies (Devlin et al., 1991). Kopp and 

colleagues (1999) attempted to document inflammation of the upper airways in response to 

summer season O3 exposures by following a group of 170 school children in two towns in the 

German Black Forest from March to October of 1994. To assess inflammation, nasal lavage 

samples were collected at 11 time points spanning the follow-up period. The mean 1⁄2-h max O3 

7-55
concentrations were 33 ppb (5th % to 95th %: 1 to 72) in Villingen and 54 ppb (5th % to 

95th %: 23 to 92) in Freudenstadt. The nasal lavage samples were analyzed for markers of 

inflammation, including eosinophil cationic protein, albumin, and leukocyte counts. Subjects 

who were sensitized to inhaled allergens were excluded. When analyzed across the entire 

follow-up period, no association was detected between upper airway inflammation and O3 

concentrations. More detailed analysis showed that the first significant O3 episode of the 

summer was followed by a rise in eosinophil cationic protein levels; however, subsequent 

and even higher O3 episodes had no effect. These findings suggest an adaptive response of 

inflammation in the nasal airways that is consistent with controlled human exposure studies (see 

Section 6.9, Effects of Inflammation and Host Defense). 

Frischer and colleagues (1993) collected nasal lavage samples from 44 school children in 

Umkirch, Germany the morning after “low” O3 days (<140 μg/m3

 or approximately 72 ppb) and 

“high” O3 days (>180 μg/m3

 or approximately 93 ppb) to measure levels of biochemical markers 

of inflammation. The researchers found that higher O3 levels were associated with increased 

polymorphonuclear leukocyte counts in all children, and increases in myeloperoxydases and 

eosinophilic cation proteins among children without symptoms of rhinitis (n = 30). These results 

indicated that O3 was associated with inflammation in the upper airways. Frischer et al. (1997) 

further investigated whether hydroxyl radical attacks played a role in mediating the O3­

associated inflammatory response of the airways. Ortho- and para-tyrosine levels were 

measured in the nasal lavage samples and the ortho/para radical ratio was used to determine the 

generation of hydroxyl radicals. Significant increases in the ortho/para ratio were observed on 

days following high ambient O3 levels. However, the ortho/para ratio was not related to 

polymorphonuclear leukocyte counts, suggesting that there was no detectable relationship 

between hydroxyl radical attacks and the inflammatory response seen in these children. Similar 

to the study by Kopp et al. (1999), the ortho/para ratio decreased at the end of the summer, 

although O3 concentrations were still high—thus providing additional evidence for a possible 

adaptive response. These findings, however, do not preclude the possibility that other 

unmeasured effects, including cell damage or lower airway responses, may have occurred with 

ongoing summer season exposures. In fact, a study of joggers repeatedly exposed to O3 while 

exercising over the summer in New York City suggested that cell damage may occur in the 

absence of ongoing inflammation (Kinney et al., 1996b). 

7-56
FEF

In two Mexico City studies by Romieu et al. (1998, 2002), the effect of antioxidant 

supplements on the association between O3 and lung function in outdoor workers and asthmatic 

children was investigated. Mean 1-h max O3 concentrations were 123 ppb for the outdoor 

workers and 102 ppb for the children. Romieu and colleagues (1998) observed significant 

inverse associations between O3 and lung function parameters, including FVC, FEV1, and 

25-75 (forced expiratory flow at 25 to 75% of FVC), among outdoor workers who received the 

placebo but not among those taking the antioxidant supplement during the first phase of testing. 

Likewise, O3 concentrations were associated with declines in lung function among children with 

moderate-to-severe asthma who were on the placebo, but no associations were found among 

those who were taking the vitamin C and E supplement (Romieu et al., 2002). These results 

indicate that supplementation with antioxidants may modulate the impact of O3 exposure on the 

small airways of two potentially at-risk populations, i.e., outdoor workers and children with 

moderate-to-severe asthma. In a further analysis of the study conducted in asthmatic children, 

genetic factors were found to contribute to the variability between individuals in the effects of O3 

on lung function (Romieu et al., 2004). Individuals with polymorphism of the glutathione 

S-transferase gene (GSTM1 null genotype) lack glutathione transferase enzyme activity, which 

plays an important role in protecting cells against oxidative damage. Results from this analysis 

indicate that asthmatic children with GSTM1 null genotype were found to be more susceptible to 

the impact of O3 exposure on small airways function. Romieu et al. (2004) noted that 

supplementation with the antioxidant vitamins C and E above the minimum daily requirement 

might compensate for the genetic susceptibility. 

7.2.6 Acute Ozone Exposure and School Absences 

The association between school absenteeism and ambient air pollution was assessed in a 

limited number of studies (Chen et al., 2000; Gilliland et al., 2001). In the study by Chen and 

colleagues (2000), daily school absenteeism was examined in 27,793 students (kindergarten to 

sixth grade) from 57 elementary school students in Washoe County, NV over a 2-year period. 

The mean 1-h max O3 was 37.5 ppb (SD 13.4) during the study period. One major limitation of 

this study was that the percent of total daily absences was the outcome of interest, not illnessrelated absences, because reasons for absences were not noted in all schools. In models 

adjusting for PM10 and CO, ambient O3 levels were associated with school absenteeism. With a 

7-57
distributed lag of 1 to 14 days, O3 concentrations were associated with a 10.4% (95% CI: 2.7, 

18.1) excess rate of school absences per 40 ppb increase in 1-h max O3. PM10 and CO 

concentrations also were associated with school absenteeism; but, the effect estimate for PM10 

was negative. The inverse relationship between O3 and PM10 may have partially attributed to the 

negative association observed between PM10 and school absenteeism. 

Ozone-related school absences also were examined in a study of 1,933 fourth grade 

students from 12 southern California communities participating in the Children’s Health Study 

(Gilliland et al., 2001). Due to its comprehensive characterization of health outcomes, this study 

is valuable in assessing the effect of O3 on illness-related school absenteeism in children. The 

study spanned a period, January through June 1996, that captured a wide range of exposures 

while staying mostly below the highest levels observed in the summer season. Mean 8-h avg O3 

(10 a.m.-6 p.m.) concentrations ranged from 35 to 55 ppb across the 12 communities. All school 

absences that occurred during this period were followed up with phone calls to parents to 

determine whether they were illness-related. For illness-related absences, further questions 

assessed whether the illness was respiratory or gastrointestinal, with respiratory symptoms 

including runny nose/sneeze, sore throat, cough, earache, wheezing, or asthma attacks. Multiple 

pollutants were measured at a central site in each of the 12 communities. A two-stage GAM was 

used to examine the effects of O3 on school absences. The analysis controlled for temporal 

cycles, day of week, and temperature, and expressed exposure as a distributed lag out to 30 days. 

The 30-day distributed lag was found to best fit the data. Associations were found between the 

30-day distributed lag of 8-h avg O3 (10 a.m.-6 p.m.) and all illness-related absence categories. 

A 108% (95% CI: 29, 235) increase in illness-related absences was observed with a 30 ppb 

increase in 8-h avg O3. Larger O3 effects were seen for respiratory causes (147% [95% CI: 6, 

478] than for nonrespiratory causes (61% [95% CI: 9, 138]). Among the respiratory absences, 

larger effects were seen for lower respiratory diseases than for upper respiratory diseases. 

Multipollutant analyses were not performed; however, in single-pollutant models neither PM10 

or NO2 were associated with any respiratory or nonrespiratory illness-related absences. Some 

concern exists regarding the possibility of residual seasonal confounding given the 6-month time 

span of the monitoring period and the long lag periods of exposure, which are likely to capture 

seasonally changing factors such as pollen episodes. Further, the biological relevance of O3 

concentrations lagged 30 days present an interpretive challenge. 

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The school absenteeism data from the Children’s Health Study was used to illustrate 

alternative new modeling approaches in two subsequent papers. Berhane and Thomas (2002) 

noted that systematic comparison of modeling approaches could potentially provide more insight 

into the modeling process. They used a two-stage iteratively weighted filtered least squares 

model which allowed explicit modeling of the autoregressive structure and overdispersion; this 

model provided information pertaining to how air pollution operates in various communities. 

Results indicated that the overall model might not be adequate for some communities, but the 

community-specific correction for overdisperion and autocorrelation compensated for this 

shortcoming. For a 30-day distributed lag, the effect estimate for illness-related absences was 

similar, 105% (95% CI: !10, 368) increase per 30 ppb increase in 8-h avg O3. The largest O3 

effect was observed in the communities with low long-term O3, PM10, and NO2 concentrations. 

Once again, PM10 was not associated with illness-related absences. Analyses for specific illness 

categories were not conducted. Berhane and Thomas (2002) also examined effects from longterm exposure to air pollutants. Long-term average concentrations of PM10, but not O3, were 

associated with increases in illness-related absenteeism. 

A further analysis by these authors (Rondeau et al., 2005) used another new approach, 

a three-stage logistic transition model. Unlike the previous two analyses, this analysis allowed 

the simultaneous examination of the effects of daily exposure to air pollution and individual risk 

factors using binary time-series data structures, without aggregating over subjects or time. 

In contrast to the results from Berhane and Thomas (2002), the Rondeau et al. results suggested 

a chronic effect of O3 on school absenteeism (41% [95% CI: 2, 95] increase per 30 ppb increase 

in 8-h avg O3 over a 5-day lag period) but no acute effect (!0.2% [95% CI: !24, 31] with a 

30-day distributed lag) after adjustment for individual factors. The acute O3 effect on respiratory 

absences was positive but nonsignificant (12% [95% CI: !22, 62]). Rondeau et al. did not 

compare their results to those from Gilliland et al. Statistical tests conducted to compare the 

effect estimates indicated that the acute O3 estimates for total illness-related absences were 

significantly different (p = 0.03) in the two studies; however, the estimates for respiratory 

absences were not found to be different (p = 0.15). Both acute and chronic exposure to PM10 

was not associated with illness-related school absences. The authors noted that the different 

results compared to the analyses by Gilliland et al. (2001) and Berhane and Thomas (2002) 

might be due to different modeling approaches or the use of binary instead of Poisson time-series 

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data structures. The two methods papers by Berhane and Thomas (2002) and by Rondeau et al. 

(2005) focused on introducing new methods and used the Children’s Health Study data for 

illustrative purposes, as opposed to comprehensive analysis of the substantive findings. 

Hence, greatest emphasis should be placed on the results from Gilliland et al (2001). 

Limited evidence from Chen et al. (2000) and Gilliland et al. (2001) suggest that 

ambient O3 concentrations, accumulated over 2 to 4 weeks, may be associated with school 

absenteeism, particularly illness-related absences. Further replication is needed before firm 

conclusions can be reached regarding the effect of O3 on school absences. 

7.2.7 Cardiovascular Endpoints 

Several air pollution studies have examined various cardiovascular endpoints (see 

Table AX7-2 in Annex 7, Section AX7.1). The earlier studies focused on PM effects. For a 

more thorough discussion of these PM studies and their health endpoints, refer to Section 8.3.1 

of the 2004 PM AQCD (U.S. Environmental Protection Agency, 2004). More recent studies 

have examined associations of O3 and other pollutants with various measures of heart beat 

rhythms in panels of elderly subjects, as discussed below. Other studies examined the increased 

risk of MI related to air pollutant exposures. 

7.2.7.1 Cardiac Autonomic Control 

Alterations in heart rate and/or rhythm are thought to reflect pathophysiologic changes that 

may represent possible mechanisms by which ambient air pollutants such as O3 may exert acute 

effects on human health. Decreased HRV has been identified as a predictor of increased 

cardiovascular morbidity and mortality. Brook et al. (2004) state that HRV, resting heart rate, 

and blood pressure are modulated by a balance between the two determinants of autonomic tone 

(the sympathetic and parasympathetic nervous systems). They noted that decreased HRV 

predicts an increased risk of cardiovascular morbidity and mortality in the elderly and those with 

significant heart disease, which is generally determined by analyses of time (e.g., standard 

deviation of normal R-R intervals) and frequency domains (e.g., low frequency/high frequency 

ratio by power spectral analysis, reflecting autonomic balance) measured during 24 hours of 

electrocardiography. Decreased parasympathetic input to the heart may provide an important 

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mechanistic link between air pollution and cardiovascular mortality by promoting fatal 

tachyarrhythmias. 

The potentially adverse effects of air pollutants on cardiac autonomic control were 

examined in a large population-based study, among the first in this field. Liao et al. (2004) 

investigated short-term associations between ambient pollutants and cardiac autonomic control 

from the fourth cohort examination (1996-1998) of the population-based Atherosclerosis Risk in 

Communities Study (ARIC). PM10, O3, and other gaseous air pollutants were examined in this 

study. PM10 (24-h avg) and O3 exposures (8-h avg, 10 a.m. to 6 p.m.) one day prior to the 

randomly allocated examination date were used. The mean 8-h avg O3 level was 41 ppb 

(SD 16). They calculated 5-minute HRV indices between 8:30 a.m. and 12:30 p.m. and used 

logarithmically-transformed data on high-frequency (0.15 to 0.40 Hz) and low-frequency 

(0.04 to 0.15 Hz) power, standard deviation of normal R-R intervals, and mean heart rate. The 

effective sample sizes for O3 and PM10 were 5,431 and 4,899, respectively, from three U.S. study 

centers in North Carolina, Minnesota, and Mississippi. PM10 concentrations measured one 

day prior to the HRV measurements were inversely associated with both frequency- and 

time-domain HRV indices. Ambient O3 concentrations were inversely associated with 

high-frequency power among whites. Consistently more pronounced associations were 

suggested between PM10 and HRV among persons with a history of hypertension. Liao et al. 

(2004) noted that these findings may represent potentially important arrhythmogenic 

mechanisms of ambient air pollution. The acute adverse effect of air pollution on cardiac 

autonomic control is based on the hypothesis that increased air pollution levels may stimulate the 

autonomic nervous system and lead to an imbalance of cardiac autonomic control characterized 

by sympathetic activation unopposed by parasympathetic control. Such an imbalance of cardiac 

autonomic control may predispose susceptible people to greater risk of life-threatening 

arrhythmias and acute cardiac events. The Liao et al. (2004) findings were cross-sectionally 

derived from a population-based sample and reflect the short-term effects of air pollution on 

HRV. When the regression coefficients for each individual pollutant model were compared, the 

effects for PM10 were considerably larger than the effects for gaseous pollutants such as O3. 

Because of the population-based sample, this study does have better generalizability than other 

smaller panel studies. The findings are suggestive of short-term effects of air pollutants, 

including O3, on HRV at the population level. 

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Another population-based study of air pollutants and HRV was conducted in Boston, MA 

on 497 men from the VA Normative Aging Study (NAS) (Park et al., 2005). The mean 24-h 

avg O3 concentration was 23.0 ppb (SD 13.0). Several associations with HRV outcomes were 

observed with a 4-h moving average of O3 concentrations; effects generally diminished with 

longer averaging times of 24 and 48 hours. Stronger associations were reported with PM2.5. 

In two-pollutant models, the magnitude of the percent changes for both PM2.5 and O3 diminished 

slightly. In analyses by ischemic heart disease, hypertension, and diabetes status, stronger 

associations of HRV with O3 and PM2.5 were observed for individuals with ischemic heart 

disease and hypertension. These results are consistent with a Mexico City study (n = 34) by 

Holguín et al. (2003) which reported an HRV effect for O3 in subjects with hypertension. The 

association of O3 exposure with reduced low-frequency power in the full cohort seemed to be 

driven by subjects not taking calcium-channel blockers (Park et al., 2005). This suggests that 

this drug is blocking effects of O3 on the sympathetic pathway. This study cohort consists of all 

males and almost all whites. This population-based study suggests that short-term exposure 

to O3 is a predictor of alterations in cardiac autonomic function as measured by HRV among 

older male adults. A potential limitation of this study is that electrocardiograms were only taken 

once for each subject, so subject-specific variation of HRV measures may not be ruled out as a 

potential confounder. 

Two separate analyses of the same cohort of patients examined the association between air 

pollution and the incidence of ventricular arrhythmias (Dockery et al., 2005; Rich et al., 2005). 

A total of 203 patients with implanted cardioverter defibrillators who lived within 25 miles of 

the ambient monitoring site at the Harvard School of Public Health, Boston, MA, were 

monitored. They had a total of 635 person-years of follow-up or an average of 3.1 years per 

subject. The median 48-h avg O3 concentration was 22.9 ppb (IQR 15.4). In the analysis by 

Dockery et al. (2005), positive associations were observed between ventricular arrhythmias 

within 3 days of a prior event and a 2-day mean of several air pollutants, including PM2.5, black 

carbon, NO2, CO, and SO2. No associations were observed with O3. There was, however, a 

suggestion of increasing risk with increasing quintiles of O3 (p <0.05). The analysis by Rich 

et al. (2005) observed stronger O3 effects on ventricular arrhythmias using a case-crossover 

study design. Case periods were defined by the time each arrhythmic event began; for each case, 

three to four control periods were selected by matching on weekday and hour of the day within 

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the same calender month. The median 24-h avg O3 level was 22.6 ppb (IQR 15.7). For a 20 ppb 

increase in 24-h moving average O3, a 27% (95% CI: 0, 60) increased risk of ventricular 

arrhythmias was estimated. Significant effects also were found for PM2.5, NO2, and SO2. In twopollutant models, the O3 effect was found to be generally robust. Stratified analysis by the 

presence of a recent ventricular arrhythmia within the previous three days indicated that O3 was 

associated with increased risk among subjects without a recent event (37% [95% CI: 6, 79]), but 

not among those with recent events (5% [95% CI: !27, 49]). Rich et al. (2005) explained that 

the use of the case-crossover study design and conditional analysis might have contributed to the 

stronger associations observed in their study compared to Dockery et al. (2005). The 

case-crossover design and conditional analysis controlled for season, time trends, weekday, as 

well as other non-time-varying confounders such as underlying medical conditions and smoking 

status by design, thereby eliminating any residual confounding. In the analysis by Dockery et al. 

(2005), these factors were controlled through modeling. In addition, the use of a 24-h moving 

average instead of a calendar-day air pollution concentration might have reduced exposure 

misclassification, resulting in larger effect estimates. 

Other studies do not provide evidence for an O3 effect on HRV and cardiac arrhythmias 

(Peters et al., 2000a; Rich et al., 2004; Vedal et al., 2004). These studies, however, may have 

had limited power to examine subtle effects. Gold et al. (2000; reanalysis Gold et al., 2003) 

reported results that suggest that O3 exposure may decrease vagal tone, leading to reduced HRV. 

In this Boston, MA study, the mean 1-h max O3 level was 25.7 ppb (IQR 23.0). In another 

Boston study, Schwartz et al. (2005) reported a weak association of O3 with the root mean 

squared differences between adjacent R-R intervals in a study of 28 elderly subjects. The 

median 1-h max O3 level was 34 ppb (IQR 26). The authors noted that lack of personal exposure 

measurements might render such studies less able to assess autonomic functions. This study by 

Schwartz et al. (2005) reported the strongest effects for black carbon. 

7.2.7.2 Acute Myocardial Infarction 

The effect of O3 on the incidence of MI was examined in a limited number of studies. 

Acute MI was studied in relation to air pollution in Toulouse, France based on the existence of 

an acute MI registry (Monitoring Trends and Determinants in Cardiovascular Disease 

[MONICA]) and an air quality network covering the same population (Ruidavets et al., 2005). 

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The mean 8-h max O3 level was 38.6 ppb (range 2.0-82.6). After adjustment for temperature, 

relative humidity, and influenza epidemics, the relative risk of acute MI occurrence was 1.76 

(95% CI: 1.12, 2.45) for current-day O3 concentrations. The increased risk of MI was more 

evident in the oldest group, 55 to 64 years of age. Further, the oldest subjects without a personal 

history of ischemic heart disease were more susceptible to an acute event when O3 levels 

increased. No PM data was reported in this study. 

In a case-crossover study (n = 772) in Boston, MA, Peters et al. (2001) reported an odds 

ratio of 1.27 (95% CI: 0.87, 1.88) per 40 ppb increase in 2-h avg O3 (1 hour before onset of 

event). The mean 24-h avg O3 level was 19.9 ppb (5th % to 95th %: 6 to 36). Stronger effects 

on the incidence of MI were observed for PM2.5 and PM10. It should be noted that Peters et al. 

(2001) used unidirectional sampling (i.e., control periods were selected only before the case 

period) in their case-crossover design. Unidirectional sampling can lead to time trend bias and 

overlap bias (i.e., biased conditional logistic regression estimating equations), which can result 

in overestimated effects of exposure (Janes et al., 2005). 

7.2.7.3 Cardiovascular Endpoints in Human Clinical Studies 

In a controlled human exposure study discussed in Chapter 6, Sections 6.3.4 and 6.10, 

Gong et al. (1998a) studied 10 nonmedicated hypertensive and 6 healthy male adults exposed to 

0.3 ppm O3 with intermittent exercise in relation to various cardiovascular effects. The overall 

results did not indicate acute cardiovascular effects of O3 in either the hypertensive or control 

subjects. The authors observed an increase in rate-pressure product and heart rate, a decrement 

for FEV1, and a >10 mm Hg increase in the alveolar/arterial pressure difference for O2 

following O3 exposure. These findings suggest that O3 may exert cardiovascular effects 

indirectly by impairing alveolar-arterial O2 transfer and potentially reducing O2 supply to the 

myocardium. Ozone exposure may increase myocardial work and impair pulmonary gas 

exchange to a degree that could perhaps be clinically important in persons with significant preexisting cardiovascular impairment. 

7.2.7.4 Summary of Field Studies with Cardiovascular Outcomes 

A limited epidemiologic database examining cardiovascular outcomes in relation to O3 

exposures is available. Among these studies, three were population-based and involved cohorts 

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such as the ARIC (Liao et al., 2004), MONICA (Ruidavets et al., 2005), and NAS (Park et al., 

2005). Such studies may offer more informative results based on their large subject-pool and 

design. Results from these three studies were suggestive of an association between O3 exposure 

and the cardiovascular endpoints studied. As in the case of respiratory disease outcomes, Brook 

et al. (2004) stated that the increase in relative risk for cardiovascular disease due to air pollution 

is small compared with the impact of the established cardiovascular risk factors. However, 

because of the enormous number of people affected, even conservative risk estimates can 

translate into a substantial increase in mortality due to cardiovascular disease within the 

population and, therefore, could potentially imply a notable public health problem. 

7.2.8 Summary of Field Studies Assessing Acute Ozone Effects 

C Results from recent field/panel studies support the evidence from controlled human 

exposure studies that acute O3 exposure is associated with a significant effect on lung 

function, as indicated by decrements in FEV1, FVC, and PEF. The declines in lung 

function were noted particularly in children and asthmatics. 

C Limited evidence suggests that more time spent outdoors, higher levels of exertion, 

and the related increase in O3 exposure may potentiate the risk of respiratory effects. 

In addition to children and asthmatics, adults who work or exercise outdoors may be 

particularly vulnerable to O3-associated health effects. 

C Many new studies have examined the association between O3 concentrations and a 

wide variety of respiratory symptoms (e.g., cough, wheeze, production of phlegm, 

and shortness of breath). Collectively, the results suggest that acute exposure to O3 

is associated with increased respiratory symptoms and increased as-needed medication 

use in asthmatic children. 

C Other panel studies evaluated O3 effects on other health outcomes, including school 

absences and markers of inflammation and oxidative damage. Ozone exposure was 

associated with increased inflammation and generation of hydroxyl radicals in the 

upper airways. Use of antioxidant supplements was found to diminish the O3 effect on 

lung function. Some studies suggest that O3 exposure, accumulated over 2-4 weeks, is 

associated with increases in respiratory-related school absences, but further replication 

is needed before firm conclusions can be drawn about O3 effects on school absences. 

C Some field studies have examined the association between O3 and cardiac physiologic 

outcomes. The current evidence is rather limited but suggestive of a potential effect on 

HRV, ventricular arrhythmias, and MI incidence. 

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7.3 ACUTE EFFECTS OF OZONE ON DAILY EMERGENCY 

DEPARTMENT VISITS AND HOSPITAL ADMISSIONS 

7.3.1 Summary of Key Findings on Studies of Emergency Department Visits 

and Hospital Admissions from the 1996 Ozone AQCD 

In the 1996 O3 AQCD, aggregate population time-series studies of O3-related health effects 

provided relevant evidence of acute responses, even below a 1-h max O3 of 0.12 ppm. 

Emergency room visits and hospital admissions were examined as possible outcomes following 

exposure to O3. In the case of emergency room visits, the evidence was limited (Bates et al., 

1990; Cody et al., 1992; Weisel et al., 1995; White et al., 1994), but results generally indicated 

an O3 effect on morbidity. The strongest and most consistent evidence of O3 effects, at levels 

both above and below 1-h max O3 levels of 0.12 ppm, was provided by the multiple studies that 

had been conducted on summertime daily hospital admissions for respiratory causes in various 

locales in eastern North America (Bates and Sizto, 1983, 1987, 1989; Burnett et al., 1994; 

Lipfert and Hammerstrom, 1992; Thurston et al., 1992, 1994). These studies consistently 

demonstrated that O3 air pollution was associated with increased hospital admissions, accounting 

for roughly one to three excess respiratory hospital admissions per million persons with each 

100 ppb increase in 1-h max O3. This association had been shown to remain even after 

statistically controlling for the possible confounding effects of temperature and copollutants 

(e.g., H+

, SO4 

!2

, PM10), as well as when considering only days with 1-h max O3 concentrations 

below 0.12 ppm. Overall, the aggregate population time-series studies considered in the 1996 O3 

AQCD provided strong evidence that ambient exposures to O3 can cause significant 

exacerbations of preexisting respiratory disease in the general public. 

7.3.2 Review of Recent Studies of Emergency Department Visits for 

Respiratory Diseases 

Emergency department visits represent an important acute outcome that may be affected 

by O3 exposures. Morbidities that result in emergency department visits are closely related to, 

but are generally less severe than, those that result in unscheduled hospital admissions. In many 

cases, acute health problems are successfully treated in the emergency department; however, 

a subset of more severe cases that present initially to the emergency department may require 

hospital admission. 

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Several studies have been published in the past decade that examined temporal associations 

between O3 exposures and emergency department visits for respiratory diseases (Table AX7-3 in 

Annex 7, Section AX7.1). Total respiratory causes for emergency room visits typically include 

asthma, pneumonia, bronchitis, emphysema, upper and lower respiratory infections such as 

influenza, and a few other minor categories. Asthma visits typically dominate the daily 

incidence counts. Chronic bronchitis and emphysema often are combined to define COPD, 

which is a prominent diagnosis among older adults with lung disease. Figure 7-8 presents 

percent changes in emergency department visits for asthma from single-pollutant models, 

with results expressed in standardized increments. The lags presented in the figure vary 

depending on reported results. Most studies reported effect estimates from a short lag period 

(0 to 2 days). Results from Weisel et al. (2002) are not included as comparable risks estimates 

for O3 were not presented. Several of the studies conducted in the United States and Canada 

examined the effects of O3 during the warm season only. In general, ambient O3 concentrations 

were associated with emergency department visits for asthma in warm-season only analyses. 

The warm-season effect estimates tended to be positive and larger than results for cool-season or 

all-year analyses. 

Among studies with adequate controls for seasonal patterns, many reported at least one 

positive association with O3. These studies examined emergency department visits for total 

respiratory complaints (e.g., Delfino et al., 1997b, 1998b; Herñandez-Gardûno et al., 1997; 

Ilabaca et al., 1999; Jones et al., 1995; Lin et al., 1999), asthma (e.g., Friedman et al., 2001; 

Jaffe et al., 2003; Stieb et al., 1996; Tenías et al., 1998; Tobías et al., 1999; Tolbert et al., 2000; 

Weisel et al., 2002), and COPD (Tenías et al., 2002). 

One recent study examined emergency department visits for total and cause-specific 

respiratory diseases in Atlanta, GA over an 8-year period (Peel et al., 2005). A distributed lag of 

0 to 2 days was specified a priori. The mean 8-h max O3 concentration was 55.6 ppb (SD 23.8). 

Ozone concentrations were associated with emergency department visits for total respiratory 

diseases and upper respiratory infections in all ages. A marginally significant association was 

observed with asthma visits (2.6% [95% CI: !0.5, 5.9] excess risk per 30 ppb increase in 8-h 

max O3), which became stronger when the analysis was restricted to the warm months (3.1% 

[95% CI: 0.2, 6.2] excess risk). In multipollutant models adjusting for PM10, NO2 and CO, O3 

was the only pollutant that remained significantly associated with upper respiratory infections. 

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7-68 

Figure 7-8. Ozone-associated percent change (95% CI) in emergency department visits for asthm

standardized increment (see Section 7.1.3.2). 

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Another large asthma emergency department study was carried out during the months of May 

through September from 1984 to 1992 in St. John, New Brunswick, Canada (Stieb et al., 1996). 

The mean 1-h max O3 level was 41.6 ppb (range 0-160). Effects were examined separately 

among children aged less than 15 years and in persons aged 15 years and older. A significant 

effect of O3 on emergency department visits was reported among persons 15 years and older. 

There was suggestion of a threshold somewhere in the range below a 1-h max O3 of 75 ppb. 

Another study in Valencia, Spain from 1994 to 1995 found that emergency room visits for 

asthma among persons over 14 years old were robustly associated with relatively low O3 levels 

(mean 1-h max O3 of 32.4 ppb [range 6.9-81.2]) (Tenías et al., 1998). The excess risk of asthma 

emergency room visits was larger in the warm season (May to October; mean 1-h max O3 of 

38.2 ppb), with an 85% (95% CI: 20, 188) excess risk per 40 ppb increase in 1-h max O3. 

During the cool season (November-April; mean 1-h max O3 of 26.5 ppb), the excess risk was 

31% (95% CI: !24, 125). 

Among the studies that reported a positive association between O3 and emergency 

department visits for respiratory outcomes, O3 effects were found to be robust to adjustment 

for PM10, NO2, SO2, and BS (Lin et al., 1999; Peel et al., 2005; Tenías et al., 1998). One study 

by Tolbert and colleagues (2000) observed that the significant univariate effects of both O3 

and PM10 on pediatric asthma emergency department visits in Atlanta, GA became 

nonsignificant in two-pollutant regressions, reflecting the high correlation between the two 

pollutants (r = 0.75). 

For several other studies with total respiratory and asthma outcomes, inconsistencies 

confound the interpretation of likely effects. For example, in a Montreal, Canada study by 

Delfino et al. (1997b), O3 effects on total respiratory emergency department visits were seen in a 

short data series from the summer of 1993 (mean 8-h max O3 of 30.7 ppb [SD 11.5]) but not in a 

similar data series from the summer of 1992 (mean 8-h max O3 of 28.8 ppb [SD 11.3]). The 

significant 1993 results were seen only for persons older than 64 years. A very similar analysis 

of two additional summers (1989 [mean 8-h max O3 of 44.1 ppb] and 1990 [mean 8-h max O3 

of 35.4 ppb]) revealed an O3 association only for 1989, but again only in persons over 64 years 

old (Delfino et al., 1998b). An analysis of data on respiratory emergency department visits 

from June to August of 1990 (mean 24-h avg O3 of 28.2 ppb [SD 11.7]) in Baton Rouge, LA 

reported O3 effects in adults, but not in children or among the elderly (Jones et al., 1995). 

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Tobías and colleagues (1999) showed that regression results for asthma emergency 

department visits could be quite sensitive to methods used to control for asthma epidemics. 

Ozone was associated with the outcome variable in only one of eight models tested. An Atlanta, 

GA study by Zhu et al. (2003) examined asthma emergency department visits of children during 

three summers using Bayesian hierarchical modeling to address model variability. Data were 

analyzed at the ZIP code level to account for spatially misaligned longitudinal data. Results 

indicated a positive, but nonsignificant relationship between O3 and emergency room visits 

for asthma. Ozone levels were not reported in this study. 

Other studies also reported no association between O3 and emergency department visits for 

respiratory causes (Atkinson et al., 1999a; Castellsague et al., 1995; Chew et al., 1999; Hwang 

and Chan, 2002; Sunyer et al., 1997). Using Bayesian hierarchical modeling, Hwang and Chan 

(2002) examined the effect of air pollutants on daily clinic visits for lower respiratory illnesses 

across 50 communities in Taiwan. The mean 1-h max O3 for all 50 communities was 54.2 ppb, 

with individual-community means ranging between 38.9 to 78.3 ppb. All pollutants except O3 

were associated with daily clinic visits. In a pooled analysis of emergency admissions for 

asthma in four European cities as part of the Air Pollution on Health: European Approach 

(APHEA) study, no overall effect of O3 was observed (Sunyer et al., 1997). Median 1-h max O3 

levels were relatively low, ranging from 14 to 37 ppb across the four cities. Atkinson et al. 

(1999a) also did not find an association in London, England between O3 and emergency 

department visits at a mean 8-h max O3 of 17.5 ppb (SD 11.5). One study by Thompson et al. 

(2001) in Belfast, Northern Ireland found no O3 effect in the warm season, but a decreased risk 

of childhood asthma admissions (!21% [95% CI: !33, !6] per 20 ppb increase in 24-h avg O3) 

in the cold season. The O3 levels were similar in both seasons, with mean 24-h avg O3 

concentrations of 18.7 ppb in the warm season and 17.1 ppb in the cold season. After adjusting 

for benzene levels, O3 was no longer associated with asthma emergency department visits during 

the cold season. The inverse relationship of O3 with benzene concentrations (r = !0.65), and 

perhaps with other pollutants, might have contributed to the apparent protective effect of O3. 

A study by Hajat et al. (1999, 2002) of physician consultations for asthma, lower respiratory 

diseases, and upper respiratory diseases in London reported negative associations with O3, which 

was also suggestive of residual confounding by copollutants or weather factors (note that data 

were analyzed using Poisson GAM with default convergence criteria). Several other emergency 

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department studies looking at O3 are more difficult to interpret because of inadequate control for 

seasonal patterns, very low O3 levels, or because no quantitative results were shown for O3 

(Buchdahl et al., 1996, 2000; Garty et al., 1998; Holmén et al., 1997; Lierl and Hornung, 2003; 

Lipsett et al., 1997; Nutman et al., 1998). 

Although several studies found a significant association between O3 concentrations and 

emergency department visits for respiratory causes, some inconsistencies were observed. The 

inconsistencies may be attributable, at least partially, to differences in model specifications and 

analysis approach among the various studies. For example, ambient O3 concentrations, length of 

the study period, and statistical methods used to control confounding by seasonal patterns and 

copollutants appear to affect the observed O3 effect on emergency department visits. Studies 

that stratified analyses by season generally reported a positive association between O3 

concentrations and emergency department visits for asthma in the warm season. 

7.3.3 Studies of Hospital Admissions for Respiratory Diseases 

Hospital admissions represent a medical response to a serious degree of morbidity for a 

particular disease. Scheduled hospitalizations are planned in advance when a particular clinical 

treatment is needed. However, unscheduled admissions are ones that occur in response to 

unanticipated disease exacerbations and are more likely to be affected by environmental factors, 

such as air pollution. As such, the hospital admissions studies reviewed here focused 

specifically on unscheduled admissions. Study details and results from hospital admissions 

studies published during the past decade are summarized in Table AX7-4 (in Annex 7, 

Section AX7.1). As a group, these hospitalization studies tend to be larger, in terms of 

geographic and temporal coverage, and indicate results that are generally more consistent than 

those reviewed above for emergency department visits. As in the case for all studies that 

examine changes in aggregate measures of acute disease outcomes over time, the following 

should be considered in comparing results: (1) difference in types of respiratory diseases for 

hospital admission; (2) age of study population; (3) mean level of O3 during study; (4) singlecity versus multicity studies; (5) length of study (e.g., <5 years versus >5 years); (6) analysis by 

season versus all year; (7) O3-only versus multipollutant models; (8) number of exposure lag 

days; and (9) type of study (e.g., case-crossover versus time-series). These factors are 

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considered in the sections below with further discussion on potential confounding of the O3 

effect estimate by seasonal factors and copollutants. 

7.3.3.1 All-year and Seasonal Effects of Ozone on Respiratory Hospitalizations 

The effect of O3 on respiratory hospitalizations was examined in various studies conducted 

in the United States and abroad. Figures 7-9 and 7-10 present risk estimates from all total 

respiratory hospital admission studies. Burnett et al. (1995), which did not present quantitative 

results for O3, and Yang et al. (2003), which only presented odds ratios, were not included in the 

figures. In cases where multiple lags were presented, the multiday lag was selected to represent 

the cumulative effect from all days examined. If only single-day lags were analyzed, the effect 

estimate of the shortest lag time, usually a lag of 0 or 1 day, was presented. Figure 7-9 plots the 

effect estimates and 95% CIs from 15 studies that analyzed all-year data. The risk estimates are 

arranged by age groups. The preponderance of positive risk estimates, with some that are 

statistically significant, is readily apparent. Figure 7-10 presents the season-stratified effect 

estimates by region. For studies that reported risk estimates from all four seasons, only the 

summer and winter estimates are presented. It appears that the warm-season estimates, 

collectively, tend to be larger, positive values compared to all-year and cool-season estimates. 

All of the negative estimates were from analyses using cool-season data only, which might 

reflect the inverse correlation between O3 and one or another copollutant, especially PM, during 

that season. None of the studies presented results for the O3 effect during the cool season after 

adjusting for potential confounding by PM. 

Among the respiratory hospitalization studies, the most robust and informative results were 

observed when a broad geographic area was examined using a consistent analytical methodology 

(Anderson et al., 1997; Burnett et al., 1995, 1997a). These studies have all reported an O3 effect 

on respiratory hospital admissions. The largest such study to date was carried out using data on 

all-age respiratory hospital admissions from 16 Canadian cities with populations exceeding 

100,000 during the period 1981 to 1991 (Burnett et al., 1997a). In addition to O3, the authors 

evaluated health effects of SO2, NO2, CO, and coefficient of haze (a surrogate for black carbon 

particle concentrations). Pooling the 16 cities, a positive association was observed between 

respiratory hospital admissions and the 1-day lag O3 concentration in the spring (5.6% [95% CI: 

1.6, 9.9] excess risk per 40 ppb increase in 1-h max O3) and summer (6.7% [95% CI: 3.5, 10.0]). 

7-72
Figure 7-9. Ozone-associated percent change (95% CI) in total respiratory hospitalizations 

for all-year analyses per standardized increment (see Section 7.1.3.2). Effect 

estimates are arranged by age groups. 

7-73 
Figure 7-10. Ozone-associated percent change (95% CI) in total respiratory 

hospitalizations by season per standardized increment (see Section 7.1.3.2). 

The results for fall were also positive, though of smaller magnitude (3.8% [95% CI: !0.2, 7.9]). 

There was no evidence for an O3 effect in the winter season (!0.8% [95% CI: !4.8, 3.3]). The 

mean 1-h max O3 concentrations for the spring, summer, fall, and winter were 40 ppb, 38 ppb, 

21 ppb, and 26 ppb, respectively. Control outcomes related to blood, nervous system, digestive 

system, and genitourinary system disorders were not associated with O3. 

7-74
In a previous study focused mainly on evaluating health impacts of sulfate particles, 

Burnett et al. (1995) reported results from a time-series analysis of all-age respiratory hospital 

admissions to 168 hospitals in Ontario, Canada over a 6-year period (1983 to 1988). The mean 

1-h max O3 was 36.3 ppb for the entire year, ranging from 21.9 ppb in December to 52.9 ppb in 

July. The outcome data were prefiltered to remove seasonal variations using a weighted 19-day 

moving average. The authors reported that O3 was associated with respiratory hospital 

admissions; however, no quantitative results for O3 were presented. 

Results from an analysis of five European cities indicated strong and consistent O3 effects 

on unscheduled hospital admissions for COPD (Anderson et al., 1997). The five cities examined 

(London, Paris, Amsterdam, Rotterdam, and Barcelona) were among those included in the 

multicity APHEA study. The number of years of available data varied from 5 to 13 years among 

the cities. City-specific effect estimates were pooled across cities using weighted means. 

An association with O3 was observed in full-year analyses. All-year median 1-h max O3 levels 

ranged from 19 to 40 ppb across the five cities. Season-stratified analyses indicated that the O3 

effect was larger in the warm season (median 1-h max O3 range 25-47 ppb), 4.7% (95% CI: 1.6, 

7.9) excess risk per 40 ppb increase in 1-h max O3, compared to the cool season (median 1-h 

max O3 range 10-33 ppb), 1.6% (95% CI: !3.1, 7.9) excess risk. There was no significant 

heterogeneity in O3 effects among the cities. 

Several additional studies carried out in one or two cities over a span of five or more years 

provided substantial additional evidence regarding O3 effects on respiratory hospital admissions 

(Anderson et al., 1998; Burnett et al., 1999, 2001; Moolgavkar et al., 1997; Petroeschevsky et al., 

2001; Ponce de Leon et al., 1996; Sheppard et al., 1999 [reanalysis Sheppard, 2003]; Yang et al., 

2003). Moolgavkar and colleagues (1997) reported significant and robust O3 effects on 

respiratory hospital admissions in adults 65 years and older in Minneapolis and St. Paul, MN 

(mean 24-h avg O3 of 26.2 ppb), but not in Birmingham, AL (mean 24-h avg O3 of 25.1 ppb). 

The absence of effects in the southern city may reflect less penetration of O3 into the indoor 

environment due to greater use of air conditioning, and thus less correlation between central 

site O3 monitoring and actual exposures of the urban populace. Ozone effects on all-age and 

age-stratified asthma and total respiratory hospital admissions were observed in Brisbane, 

Australia (Petroeschevsky et al., 2001). Effect sizes were found to be consistent in the warm and 

cool seasons (data not provided). Petroeschevsky et al. commented that the year-round effect 

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of O3 might reflect the relatively small degree of seasonal variation in O3 levels observed in 

Brisbane. Mean 8-h avg O3 (10 a.m.-6 p.m.) levels for the winter, spring, summer, and fall were 

16.1 ppb, 23.3 ppb, 19.9 ppb, and 16.7 ppb, respectively. The authors also noted that given the 

subtropical climate in Brisbane, characterized by warm, dry winters, perhaps the proportion of 

the population exposed to winter O3 concentrations was higher than in cities where inclement 

winter weather might force populations indoors. 

Another set of studies examined associations between O3 and respiratory-related 

hospitalizations in single cities over shorter (<5 year) time spans. Positive and significant O3 

effects were reported in Cleveland, OH (Schwartz et al., 1996); New York City (Gwynn and 

Thurston, 2001); Northern New Jersey (Weisel et al., 2002); Toronto, Canada (Burnett et al., 

1997b); Helsinki, Finland (Pönkä and Virtanen, 1996); São Paulo, Brazil (Braga et al., 1999; 

Gouveia and Fletcher, 2000a); and Hong Kong (Wong et al., 1999a). The Helsinki study by 

Pönkä and Virtanen (1996) reported significant effects of O3 on both asthma and on digestive 

disorders in a setting of very low O3 concentrations (mean 8-h max O3 of 11 ppb), which raises 

questions of plausibility. 

Less consistent effects of O3 were seen in other respiratory hospitalization studies 

(Schouten et al., 1996; Lin et al., 2003, 2004; Linn et al., 2000; Morgan et al., 1998a; Oftedal 

et al., 2003). In a study conducted in Amsterdam and Rotterdam, the Netherlands, associations 

between O3 and respiratory admissions were observed; however, results were difficult to 

interpret due to the large number of statistical tests performed (Schouten et al., 1996). In a 

California study by Neidell (2004), a negative association was observed between hospitalizations 

for asthma and naturally occurring seasonal variations in O3 within ZIP codes for children aged 

0 to 18 years. However, the O3 effect was found to be influenced by socioeconomic status. 

Among children of low socioeconomic status, O3 generally was associated with increased 

hospitalizations, with statistical significance reached in certain age groups. Neidell further stated 

that avoidance behavior on high O3 days (i.e., 1-h max O3 >200 ppb) may have attributed to the 

negative relationship observed in children of higher socioeconomic status. 

No associations between respiratory hospital admissions and O3 were seen in studies from 

Los Angeles, CA (Nauenberg and Basu, 1999); Detroit, MI (Lippman et al., 2000 [reanalysis Ito, 

2004]); Vancouver, Canada (Lin et al., 2004); London, England (Atkinson et al., 1999b); 

Edinburgh, Scotland (Prescott et al., 1998); and Drammen, Norway (Hagen et al., 2000). 

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Several of these studies were carried out in locations with low O3 levels, suggestive of a 

nonlinear concentration-response relationship. For example, the mean 1-h max O3 was 

28.02 ppb in the Vancouver, Canada study by Lin et al. (2004). The Edinburgh, Scotland study 

by Prescott et al. (1998) had a mean 24-h avg O3 of 14.5 ppb. Inadequate control of seasonal 

confounding may underlie some of the nonsignificant and negative findings. An additional 

factor likely contributing to lack of associations observed is the relatively small sample sizes 

included in some of these studies. 

For respiratory hospitalization outcomes, the largest, most significant associations with O3 

concentrations were observed when using short lag periods, in particular a 0-day lag (exposure 

on same day) and a 1-day lag (exposure on previous day). In the study of 16 Canadian cities by 

Burnett et al. (1997a), the strongest association between O3 and respiratory hospitalizations was 

found at a 1-day lag. A decline in the magnitude and significance of the effect was seen with 

increasing days lagged for O3. Anderson et al. (1997) investigated the association between O3 

and daily hospital admissions for COPD in five European cities. Lags up to 5 days were 

examined, and the largest risk estimates were found using 0- and 1-day lags. These results 

suggest that O3 has a short-term effect on respiratory hospitalizations. 

Burnett et al. (2001) investigated the association between respiratory hospitalizations 

and O3 in children less than 2 years of age (note that analyses were performed using default 

convergence criteria for Poisson GAM with a nonparametric LOESS prefilter applied to 

pollution and hospitalization data). Lags up to 5 days were examined after stratifying by season 

(Figure 7-11). The mean 1-h max O3 during the summer season was 45.2 ppb (IQR 25). In the 

summer, significant associations between O3 and daily admissions were found in several of the 

single-day lags, with the largest risk estimate of 12.5% (95% CI: 5.7, 19.7) excess risk per 

40 ppb increase in 1-h max O3 at a 1-day lag. The positive effect estimates observed at multiple 

single-day lags indicated that O3 exposure likely had an immediate effect that persisted over 

several days. Using a cumulative lag period of 0- to 4-days, the O3-related risk estimate was 

30.2% (95% CI: 18.0, 42.4). 

Weisel et al. (2002) stated that a lag period of 1 to 3 days between exposure to O3 and 

hospital admissions or emergency department visits for asthma was plausible, because it might 

take time for the disease to progress to the most serious responses following exposure. Also, 

taking medication could further delay the progression of the adverse effect. Thus, although 

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Figure 7-11. Percent changes in total respiratory hospitalizations per 40 ppb increase in 

1-h max O3 in children less than two years of age during the summer (May to 

August). Single-day lags (0-, 1-, 2-, 3-, 4-, and 5-day) are shown in the upper 

panel. The cumulative multiday lag (0- to 4-day) is shown in the lower panel. 

Source: Derived from Burnett et al. (2001). 

strongest associations are found at lags of 0 and 1 days, examining longer single-day lag periods 

or multiday lag periods may enhance the understanding of O3 effects on hospitalizations. 

In conclusion, the evidence generally supports the findings of significant and robust effects 

of O3 on various respiratory disease hospitalization outcomes. Large multicity studies, as well as 

many studies from individual cities have reported positive O3 associations with total respiratory 

hospitalizations, asthma, and COPD, especially in studies analyzing the O3 effect during the 

summer or warm season. 

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7.3.3.2 Potential Confounding of the Ozone Effect on Respiratory Hospitalizations 

by Copollutants 

As in the case for most air pollution studies, potential confounding of the association 

between O3 and respiratory hospitalizations by copollutants generally was examined using 

multipollutant regression models. The changing relationship between O3 and copollutants by 

season complicates assessment of potential confounding of the O3 effect by copollutants. 

Figure 7-12 compares the risk estimates from models with and without adjustment for PM 

indices. This figure indicates that O3 risk estimates are fairly robust to PM adjustment in all-year 

and warm-season only data. Given the inverse relationship typically observed between O3 and 

PM during the cool season, the influence of PM adjustment on O3 risk estimates during that 

season is of particular interest. However, none of the hospitalization studies examined O3 risk 

estimates after adjusting for PM in cool-season only data. 

Several analyses of a large data set from Toronto, Canada spanning the years 1980 to 1994 

reported O3 effects on respiratory hospitalizations for all ages (Burnett et al., 1997b, 1999) and 

for persons less than 2 years old (Burnett et al., 2001). In the 1999 and 2001 studies, analyses 

were performed using Poisson GAM (default convergence criteria) with a nonparametric LOESS 

prefilter applied to the pollution and hospitalization data. All studies found that O3 effects were 

robust when adjusting for PM indices, whereas PM effects from single-pollutant models were 

markedly attenuated when O3 was added to the regression. These results imply more robust 

associations with respiratory hospitalizations for O3 than PM. 

Results from the APHEA study indicated strong and consistent O3 effects on unscheduled 

hospital admissions for COPD (Anderson et al., 1997). Significant effects also were seen for 

BS, TSP, and NO2. The authors reported that among all pollutants examined, the most consistent 

and significant findings were for O3. No two-pollutant model results were reported. Several 

additional studies also observed that there was no substantial difference in the O3 effect after 

adjusting for PM in the regression model (Gouveia and Fletcher, 2000a; Petroeschevsky et al., 

2001; Ponce de Leon et al., 1996). 

Collectively, these results suggest that copollutants generally do not confound the 

association between O3 and respiratory hospitalizations. Ozone risk estimates were robust to PM 

adjustment in all-year and warm-season only data. 

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Figure 7-12. Ozone-associated percent change (95% CI) in total respiratory 

hospitalizations with adjustment for PM indices per standardized 

increment (see Section 7.1.3.2). Analyses performed using all-year 

data unless noted otherwise. 

7.3.4 Association of Ozone with Hospital Admissions for 

Cardiovascular Disease 

Some hospital admissions studies have examined the association of O3 with cardiovascular 

outcomes (see Figure 7-13). Many reported negative or inconsistent associations (Ballester 

et al., 2001; Burnett et al., 1999; Fung et al., 2005; Koken et al., 2003; Linn et al., 2000; 

7-80
Figure 7-13. Ozone-associated percent change (95% CI) in total cardiovascular 

hospitalizations per standardized increment (see Section 7.1.3.2). 

Analyses include all ages unless otherwise noted. 

7-81 
Lippmann et al., 2000 [reanalysis Ito, 2004]; Mann et al., 2002; Morgan et al., 1998a; 

Petroeschevsky et al., 2001; Poloniecki et al., 1997; Prescott et al., 1998). In the Ito (2004) 

reanalysis of the Lippmann et al. (2000) study in Detroit, MI, no associations were observed 

between cause-specific cardiovascular admissions (i.e., ischemic heart disease, dysrhythmias, 

heart failure, and stroke) and O3 concentrations (mean 24-h avg O3 of 25 ppb) using all-year 

data. The study by Linn et al. (2000) examined associations between O3 concentrations and 

cardiovascular admissions during all four seasons as well as using all-year data in Los Angeles, 

CA. Mean 24-h avg O3 levels for the spring, summer, fall, and winter were 32 ppb, 36 ppb, 

13 ppb, and 14 ppb, respectively. Positive, but nonsignificant effects were observed during the 

spring and summer (0.6% [95% CI: !1.4, 2.6] and 0.2% [95% CI: !1.7, 2.2] respectively, per 

20 ppb increase in 24-h avg O3), while negative associations were observed during the fall and 

winter (!0.6% [95% CI: !3.3, 2.2] and !4.1% [95% CI: !7.1, !1.1] respectively). 

Some other studies, especially those that examined the relationship when O3 exposures 

were higher, have found robust positive associations between O3 and cardiovascular hospital 

admissions (Atkinson et al., 1999b; Burnett et al., 1997b; Chang et al., 2005; Tsai et al., 2003a; 

Wong et al., 1999a,b; Yang et al., 2004a). For example, Burnett et al. (1997b) reported a 

positive association between O3 and cardiovascular hospital admissions in Toronto, Canada in 

a summer-only analysis (mean 1-h max O3 of 41.2 ppb). The results were robust to adjustment 

for various PM indices, whereas the PM effects diminished when adjusting for gaseous 

pollutants. Other studies stratified their analysis by temperature, i.e., by warms days ($20 °C) 

versus cool days (<20 °C). Several analyses using warms days consistently produced positive 

associations (Chang et al., 2005; Tsai et al., 2003a; Yang et al., 2004a). On the other hand, in 

two studies conducted in Hong Kong, total cardiovascular (as well as circulatory, ischemic heart 

disease, and heart failure) were all reported to be significantly associated with O3 in the cool but 

not the warm season (Wong et al., 1999a,b). In Wong et al. (1999b), O3 concentrations were 

similar in both seasons, with warm-season O3 levels being slightly lower (mean 8-h avg O3 = 

16.1 ppb) than the cool season levels (mean O3 = 18.0 ppb). The authors speculated that 

differing activity patterns and home ventilation factors may have contributed to the seasonal 

differences in O3 effects. Weather in Hong Kong is mild throughout the year, but less humid and 

cloudy in the cool season. Thus, during the cool season people are more likely to open windows 

7-82
or stay outdoors, resulting in higher personal exposures even with similar ambient O3 

concentrations across seasons. 

An increasing number of hospitalization studies have examined the association between O3 

and cardiovascular admissions. Some studies, especially those that performed stratified analyses 

by seasonal or meteorological factors, have observed positive associations. However, the overall 

evidence remains inconclusive regarding the effect of O3 on cardiovascular hospitalizations. 

7.3.5 Summary of Acute Ozone Effects on Daily Emergency Department 

Visits and Hospital Admissions 

C The vast majority of emergency room visits and hospitalization studies conducted 

over the past decade have looked at effects of O3 on either total respiratory diseases 

and/or asthma. Many of these studies analyzed O3 risk estimates using year-round 

data. Given the strong seasonal variations in O3 concentrations and the changing 

relationship between O3 and other copollutants by season, inadequate adjustment for 

seasonal effects might have masked or underestimated the association between O3 

and the respiratory disease outcomes. Ozone was generally found to be associated 

with respiratory hospitalizations and asthma emergency department visits during the 

warm season but not during the cool season. 

C Several studies have examined the association between O3 and respiratory 

hospitalizations while controlling for other pollutants in the analytical model. 

In most cases, O3 effects have been reported to be robust to adjustment for 

copollutants, particularly PM. Therefore, the evidence is supportive of independent 

O3 effects on respiratory hospital admissions. 

C Some hospital admission studies examined the effect of O3 on cardiovascular 

outcomes. A few studies observed positive O3 associations, largely in the warm 

season. Overall, however, the currently available evidence is inconclusive regarding 

any association between ambient O3 exposure and cardiovascular hospitalizations. 

7.4 ACUTE EFFECTS OF OZONE ON MORTALITY 

7.4.1 Summary of Key Findings on Acute Effects of Ozone on Mortality 

from the 1996 Ozone AQCD 

A limited number of studies, most of which were from the 1950s and 1960s, had 

examined O3-mortality associations at the time of the 1996 O3 AQCD. The 1996 O3 AQCD 

considered these historical studies to be flawed because of inadequate adjustment for seasonal 

trends or temperature and the use of questionable exposure indices. There were only a few other 

7-83
time-series studies that had examined O3-mortality associations between the 1980s and mid1990s. These latter studies used more sophisticated approaches in addressing seasonal 

confounding and weather models. One of them (Shumway et al., 1988) focused on possible 

associations with long-term O3 fluctuations in Los Angeles, CA but did not examine short-term 

associations. A study that reanalyzed Los Angeles, CA data with a focus on short-term 

associations (Kinney and Özkaynak, 1991) did find that, of the PM and gaseous criteria 

pollutants, O3 (reported as total oxidants) was most strongly associated with total nonaccidental 

mortality. Then two more studies, one conducted in Detroit, MI (O3 concentrations not 

provided) (Schwartz, 1991) and the other in St. Louis, MO (mean 24-h avg O3 of 22.5 ppb) and 

Kingston-Harriman, TN (mean 24-h avg O3 of 23.0 ppb) (Dockery et al., 1992), reported that 

PM but not O3 was significantly associated with mortality. However, the 1996 O3 AQCD noted 

that, without sufficient presentation of model specifications, it was difficult to evaluate whether 

the lack of O3-mortality associations was possibly due to mis-specification of the weather model. 

In summary, because of the insufficient number of studies that examined O3-mortality 

associations and the uncertainties regarding weather model specifications, the 1996 O3 AQCD 

was unable to quantitatively assess O3-mortality excess risk estimates or, even, to provide 

qualitative assessment of the likelihood of O3-mortality associations. 

7.4.2 Introduction to Assessment of Current Ozone-Mortality Studies 

Introductory discussions of PM-mortality effects often cite historical air pollution incidents 

such as the 1952 London, England smog episode, in which thousands of deaths were attributed 

to the air pollution from coal burning. There is no counterpart “historical episode” for O3­

mortality effects. Instead, the early recognition of the adverse health effects of summer oxidant 

air pollution, mainly from Los Angeles and other major cities with a high density of 

automobiles, were based on symptoms such as eye and throat irritations. Thus, the focus of PM 

epidemiology and that of O3 epidemiology have been historically different. 

As shown in Table AX7-5 in Annex 7, Section AX7.1, the number of short-term mortality 

studies that analyzed O3 has increased markedly since the last publication of the 1996 O3 AQCD. 

The increased attention to PM-mortality associations in the early 1990s led to an increase in 

studies that also examined O3, most often as a potential confounder for PM. Although many of 

these PM studies also reported O3 estimates, they often lacked specific hypotheses regarding 

7-84
mortality effects of O3, given that the main focus of these studies was to examine PM-mortality 

associations. This is in contrast to the O3-morbidity studies, most of which were specifically 

designed to examine effects of “summer haze” and O3 (or oxidants) on respiratory and other 

symptoms, lung function, emergency department visits, etc. However, new studies with 

hypotheses developed specifically for O3 effects on mortality have become available, such as the 

large U.S. 95 communities study by Bell et al. (2004), the U.S. 14 cities study by Schwartz 

(2005), and the 23 European cities study by Gryparis et al. (2004) discussed in the next section. 

7.4.3 Single-Pollutant Model Ozone-Mortality Risk Estimates 

To facilitate a quantitative overview of the O3-mortality effect estimates and their 

corresponding uncertainties, the percent excess risks of total nonaccidental mortality calculated 

using all-year data are plotted in Figures 7-14 and 7-15. Studies that only conducted seasonal 

analyses are presented in the next section. These figures do not include studies that only 

examined cause-specific mortality. Figure 7-14 presents only those results derived from singleday lag models. Results from multiday lag models are shown in Figure 7-15. All effect 

estimates are from single-pollutant models and include all age groups unless otherwise noted. 

The majority of the estimates are positive, with a few exceptions. Of particular note, five 

multicity studies, three from the United States (Bell et al., 2004; Samet et al., 2000 [reanalysis 

Dominici et al., 2003]; Schwartz, 2005) and two from Europe (Gryparis et al., 2004; Touloumi 

et al., 1997), showed generally positive associations. 

The initial primary objective of the original NMMAPS (Samet et al., 2000; reanalysis 

Dominici et al., 2003) was to investigate the effects of PM, but the study also examined 

mortality risk estimates from gaseous pollutants in 90 U.S. cities over the period of 1987 to 

1994. Among the 90 cities, 80 monitored O3 either year-round or during the warm season. 

The study illustrated that the mortality risk estimates for O3 varied by season. The estimate 

using all available data was about half of that for summer-only data at a lag of 1-day (see 

Section 7.6.3.2 for further discussion). Bell et al. (2004) extended the original NMMAPS 

by adding six more years (from 1987 to 2000) and 15 more communities (a total of 

95 communities), and examined the effects of O3 on mortality. Due to its extensive coverage 

and its specific focus on O3- mortality effects, Bell et al. (2004) offer a more comprehensive 

analysis than the original Samet et al. (2000) analysis. 

7-85
Figure 7-14. All cause (nonaccidental) O3 excess mortality risk estimates (95% CI) 

for all year analyses per standardized increment (see Section 7.1.3.2). 

Analyses include all ages unless otherwise noted. Only results from 

single-day lag models are presented. 

7-86 
Figure 7-15. All cause (nonaccidental) O3 excess mortality risk estimates (95% CI) 

for all year analyses per standardized increment (see Section 7.1.3.2). 

Analyses include all ages unless otherwise noted. Only results from 

multiday lag models are presented. 

The results of the study by Bell et al. (2004) are discussed in detail in this document 

because of the study’s emphasis on U.S. data and the inclusion of 95 large communities across 

the country, making this mortality study the most representative one of the U.S. population. 

In addition, this study is one of the few that have focused specifically on O3 hypotheses testing 

and investigated several important issues. Among the 95 communities examined in this study, 

55 monitored O3 throughout the year and 32 monitored O3 only during the warm season 

(generally April to October). Eight additional cities switched from warm-season only to year7-87
round monitoring or year-round to warm-season only monitoring at some point during the study 

period. Figure 7-16 presents the median 24-h avg O3 concentrations for the 95 communities 

from 1987 to 2000. In the 55 communities that had all-year data available, the median 24-h avg 

O3 concentrations ranged from 14.4 ppb in Newark, NJ to 37.3 ppb in Bakersfield, CA. In the 

40 communities with warm-season only data, the median 24-h avg O3 levels ranged from 

20.4 ppb in Portland, OR to 36.2 ppb in Memphis, TN. The range of median concentrations 

from communities monitored all year overlap considerably with the range from communities 

that only monitored during the warm season. This is expected, given that communities that have 

higher O3 concentrations year-round are generally monitored throughout the year. The mean 

24-h avg O3 concentration from all available data was approximately 26 ppb for the 

95 communities. 

Within-community results were first calculated, using single-day lags of 0, 1, 2, and 

3 days, and a 7-day distributed lag in O3 exposure. The individual-community maximum 

likelihood effect estimates for all cause mortality per 20 ppb increase in 24-h avg O3 from a 

constrained 7-day distributed lag model are presented in Figure 7-17. The heterogeneity of the 

effect estimates from the individual communities is partially attributable to differences in 

pollution characteristics, the use of air conditioning, time-activity patterns, and socioeconomic 

factors. A two-stage Bayesian hierarchical model was also used to determine a national average 

effect estimate, taking into consideration community-to-community variation. Figure 7-18 

presents the Bayesian community-specific and national average O3 risk estimates for all cause 

mortality per 20 ppb increase in 24-h avg O3 with a constrained 7-day distributed lag. The 

Bayesian community-specific estimates were shrunk to the national average estimate by a factor 

that was inversely proportional to the heterogeneity of the community-specific relative rates. 

Because of the random variation, as well as the smaller sample sizes within each city, particular 

consideration is given to the Bayesian national average effect estimate. 

In the U.S. 95 communities study, the largest risk estimate for O3-mortality was obtained 

with a 0-day lag, followed by diminishing risk estimates with 1-, 2-, and 3-day lags (Figure 7-19, 

upper panel). Ozone exposure at a 0-day lag was associated with a 0.50% (95% PI: 0.24, 0.78) 

excess risk in mortality per 20 ppb increase in 24-h avg O3. The 7-day distributed lag model, 

which examined the cumulative effect from the same day and six previous days, also is shown in 

Figure 7-19 (lower panel). A cumulative excess mortality risk of 1.04% (95% PI: 0.54, 1.55) 

7-88
Figure 7-16. Median 24-h avg O3 concentrations (10th percentile to 90th percentile range) 

for 95 U.S. communities (NMMAPS) from 1987 to 2000, arranged by O3 

concentration. Results from all available data are presented. Fifty-five of 

the 95 communities had year-round data. The remaining communities 

mostly had data during the warm season only. 

Source: Derived from iHAPSS (2005). 

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Figure 7-17. Community-specific maximum likelihood estimates for the percent change 

(95% CI) in daily mortality per 20 ppb increase in 24-h avg O3 in the 

previous week using a constrained distributed lag model for 95 U.S. 

communities (NMMAPS), arranged by size of the effect estimate. Results 

from all available data are presented. Fifty-five of the 95 communities had 

year-round data, marked by open circles ("). The cities with mostly 

warm-season data are marked by closed circles (!). 

Source: Derived from Bell (2006). 

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Figure 7-18. Community-specific Bayesian estimates and national average for the percent 

change (95% PI) in daily mortality per 20 ppb increase in 24-h avg O3 in the 

previous week using a constrained distributed lag model for 95 U.S. 

communities (NMMAPS), arranged by size of the effect estimate. Results 

from all available data are presented. Fifty-five of the 95 communities had 

year-round data, marked by open circles ("). The cities with mostly warmseason data are marked by closed circles (!). 

Source: Derived from Bell et al. (2004). 

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Figure 7-19. Percent changes in all cause mortality per 20 ppb increase in 24-h avg O3 in 

all ages. Single-day lags (0-, 1-, 2-, and 3-day) are shown in the upper panel. 

The cumulative multiday lag (0- to 6-day) is shown in the lower panel. 

Source: Derived from Bell et al. (2004). 

per 20 ppb increase in 24-h avg O3 during the previous week was observed. In a related U.S. 

study of the 19 largest cities by Huang et al. (2005), the O3 estimate for the summer season was 

1.47% (95% PI: 0.54, 2.39) excess risk of cardiopulmonary mortality with current-day 

exposure. Mean 24-h avg O3 levels ranged from 18 ppb in Oakland, CA to 56 ppb in 

San Bernadino, CA. Smaller effects also were observed with 1- and 2-day lags of exposure. 

The effect estimate for the 7-day distributed lag was 2.52% (95% PI: 0.94, 4.10) excess risk of 

cardiopulmonary mortality. These findings suggest that the effect of O3 on mortality is 

immediate, but also may persist over multiple days. 

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The influence of higher O3 levels on the risk estimate also was evaluated in the U.S. 

95 communities study. When the data were restricted to days with 24-h avg O3 levels less 

than 60 ppb for the 1-day lag analysis, the national estimate did not substantially change 

(0.30% [95% PI: 0.06, 0.54] per 20 ppb increase for days with levels below 60 ppb versus 

0.36% [95% PI: 0.12, 0.61] for all days). These results suggest that the O3-mortality 

associations may occur at 24-h avg O3 levels below 60 ppb. 

Schwartz (2005) examined O3-mortality associations using data from 14 U.S. cities. 

Median 1-h max O3 levels ranged from 35.1 ppb in Chicago, IL to 60.0 ppb in Provo, UT. 

A case-crossover study design was used to compare the influence of adjustment methods for 

temperature (regression splines of temperature versus matching case and control periods by 

temperature). The risk estimate obtained by matching (0.92% [95% CI: 0.06, 1.80] per 40 ppb 

increase in 1-h max O3 at a 0-day lag) was similar to that obtained with regression splines 

(0.76% [95% CI: 0.13, 1.40]), suggesting that the O3-mortality risk estimates were not sensitive 

to these adjustment methods for temperature. 

The APHEA 1 project (Touloumi et al., 1997) reported a pooled random effects estimate of 

4.5% (95% CI: 1.6, 7.7) per 40 ppb increase in 1-h max O3 using the best single-day lag model 

results from four European cities (Athens, Barcelona, London, and Paris). Mean 1-h max O3 

levels ranged from 21.2 ppb in London to 48.4 ppb in Athens. As an extension of the four 

European cities study, researchers of the APHEA 2 project investigated the effect of O3 on total, 

cardiovascular, and respiratory mortality in 23 cities throughout Europe (Gryparis et al., 2004). 

Ozone data were available year-round in all 23 cities. A cumulative lag of 0 to 1 days was 

hypothesized a priori. A two-stage hierarchical model, which accounted for statistical variance 

and heterogeneity among cities, was used to estimate the pooled regression coefficients. 

Because of substantial heterogeneity among cities, random effects regression models were 

applied. The pooled effect estimate for the 23 European cities (0.23% [95% CI: !0.85, 1.95] 

per 40 ppb increase in 1-h max O3 for all seasons) was positive but considerably smaller 

compared to that obtained in the APHEA 1 study. Mean O3 concentrations for all-year data were 

not presented. Gryparis et al. (2004) noted that there was a considerable seasonal difference in 

the O3 effect on mortality; thus, the small effect for the all-year data might be attributable to 

inadequate adjustment for confounding by season. This seasonal effect is discussed further in 

the next section. 

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Collectively, the single-pollutant model estimates from the single- and multiple-city 

studies shown in Figures 7-14 and 7-15 suggest that an excess risk of total nonaccidental 

mortality is associated with acute O3 concentrations. Despite the different analytical approaches 

and alternative model specifications used in the various studies, overall, most of the positive 

estimates range from 0.5 to 5% excess risk in mortality per standardized increment for all singlepollutant, all-year analyses. 

7.4.4 Meta-Analyses of Ozone-Mortality Risk Estimates 

Several studies in recent years conducted meta-analyses of O3-mortality associations (Levy 

et al., 2001; Stieb et al., 2002, 2003; Thurston and Ito, 2001; World Health Organization, 2004). 

Figure 7-20 presents the combined O3 risk estimates from the various meta-analyses. Most of 

these analyses included GAM studies using default convergence criteria except for the one by 

Stieb et al. (2003), which compared effect estimates from GAM-affected studies to non-GAM 

studies. All of these meta-analyses reported fairly consistent and positive combined estimates, 

approximately 2% excess total nonaccidental mortality per standardized increment (see Section 

7.1.3.2). However, most of these studies were not analytical in design, in that they did not 

attempt to examine the source of heterogeneity, although one suggested an influence of weather 

model specification (Thurston and Ito, 2001) and another reported evidence of publication bias 

(World Health Organization, 2004) in the past literature. None of these studies address the issue 

of season-specific estimates, therefore, interpreting these combined estimates requires caution. 

Most recently, three research groups conducted independent meta-analyses of O3-mortality 

associations (Bell et al., 2005; Ito et al., 2005; Levy et al., 2005) which attempted to evaluate the 

source of heterogeneity using the most up-to-date literature database. These analyses have also 

been systematically compared and discussed (Bates, 2005; Goodman, 2005). The all-season 

combined point estimates per standardized increment from these three meta-analyses were 

remarkably consistent: 1.75% (95% PI: 1.10, 2.37), 1.6% (95% CI: 1.1, 2.0), and 1.64% (95% 

CI: 1.25, 2.03), for the Bell et al., Ito et al., and Levy et al. studies, respectively. All three 

studies also indicated that the estimates were higher in warm seasons. Each of these studies is 

briefly summarized below. Their findings related to specific issues are discussed later in the 

corresponding sections. 

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Figure 7-20. Combined all cause (nonaccidental) O3 excess mortality risk estimates 

(95% CI) from recent meta-analyses per standardized increment 

(see Section 7.1.3.2). Note that all meta-analyses, except Stieb et al. (2003), 

included studies which used Poisson GAM with default convergence criteria. 

Bell et al. (2005) conducted a meta-analysis of 144 effect estimates from 39 U.S. and 

non-U.S. studies and estimated pooled effects by lags, age groups, specific causes, and exposure 

metrics. The results were also compared with their NMMAPS results (Bell et al., 2004). 

A two-stage Bayesian hierarchical model was used to estimate the combined estimate by taking 

into account the within-city variance (the statistical uncertainty) and between-study variance 

(the heterogeneity across cities). Bell et al. (2005) concluded that the results provided strong 

evidence of a short-term association between O3 and mortality that was not sensitive to 

adjustment for PM or for model specifications (discussed in Section 7.4.6). However, they 

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suggested that, based on comparisons between the meta-analysis results and NMMAPS results, 

there was evidence of publication bias (1.75% [95% CI: 1.10, 2.37] per 20 ppb increase in 24-h 

avg O3 for meta-analysis versus 0.50% [95% CI: 0.24, 0.78] for NMMAPS 0-day lag results). 

Ito et al. (2005) both conducted a meta-analysis of 43 U.S. and non-U.S. studies and also 

analyzed data from 7 U.S. cities to further examine issues identified by their meta-analysis. 

Adjusting for PM did not substantially influence the O3-mortality effect estimates in either the 

meta-analysis or 7 U.S. cities analysis. The multicity analysis further indicated that the 

difference in the weather adjustment model could result in a 2-fold difference in risk estimates 

(e.g., 1.96% versus 0.96% in multicity combined estimates across alternative weather models for 

the O3-only, all-year case). In the meta-analysis, they found suggestive evidence of publication 

bias (a significant asymmetry in the funnel plot), but adjusting for the asymmetry reduced the 

combined estimate only slightly (from 1.6% [95% CI: 1.1, 2.0] to 1.4% [95% CI: 0.9, 1.9] per 

20 ppb increase in 24-h avg O3). The extent of potential bias implicated in this study differed 

compared to that reported by Bell et al. (2005). The source of this difference is not clear, but Ito 

et al. stated that sensitivity analyses comparing estimates from commonly used weather model 

specifications suggest that the stringent weather model used in NMMAPS may tend to yield 

smaller risk estimates than those used in other studies. 

Levy et al. (2005) analyzed 48 estimates from 28 studies from the United States, Canada, 

and Europe using an empiric Bayesian meta-regression with covariates including the relationship 

between O3 and other pollutants, proxies for the relationship between personal exposure and 

ambient concentration such as air conditioning prevalence, and statistical methods used. They 

found that the air conditioning prevalence (a greater effect in cities with less air conditioning) 

and lag time (same-day effects larger than lagged effects) were the strongest predictors of 

between-study variability. The warm-season estimates were larger than the cool-season 

estimates. The influences of copollutants were inconsistent, but a potential influence of 

summertime PM2.5 was found. 

As stated earlier, the combined O3 excess mortality risk estimates from the meta-analyses 

by Bell et al. (2005), Ito et al. (2005), and Levy et al. (2005) were all very consistent. Although 

the analyses were conducted independently, there was considerable overlap among the risk 

estimates used in the three meta-analyses; thus, the agreement in the combined risk estimates 

was not unexpected. The common findings among these three meta-analyses, aside from the 

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consistency in their combined estimates, include: (1) no difference in estimates between GAM 

studies using default versus stringent convergence criteria; (2) larger estimates in warm seasons; 

and (3) no strong indication of PM confounding. Both the Bell et al. and Levy et al. analyses 

found that the estimates at lag 0-day were larger than for longer lags. Both the Bell et al. and Ito 

et al. studies also suggested evidence of publication bias. The positive O3 effect estimates, along 

with the sensitivity analyses in these three meta-analyses provide evidence of a robust 

association between ambient O3 and mortality. The combined effect estimates from the various 

meta-analyses ranged from 1.5 to 2.5% excess risk in all-cause mortality. 

7.4.5 Seasonal Variation in Ozone-Mortality Risk Estimates 

Since the O3 seasonal cycle follows the seasonal cycle of temperature (which is inversely 

related to the mortality seasonal cycle), inadequate adjustment of temporal trends in the 

regression model could lead to negative O3-mortality risk estimates. In addition (as discussed in 

Section 7.1.3.5) low-level O3 during the winter in some cities may be negatively correlated with 

PM and other primary pollutants, resulting in negative correlations between O3 and mortality 

even in short-term relationships. Such a confounding effect by season could be substantially 

reduced by conducting season-stratified analyses. 

A fewer number of O3-mortality studies performed seasonal analyses. Figure 7-21 presents 

the studies that reported O3 risk estimates for all-cause mortality by season. For those studies 

that obtained O3 risk estimates for each of the four seasons, only summer and winter results are 

shown. The estimates for year-round data analyses, when available, also are shown for 

comparisons. In all the studies, the O3 risk estimates are larger during the warm season than the 

cool season, with the all-year estimates generally falling between the two seasonal estimates. 

In three U.S. or European multicity studies (Gryparis et al., 2004; Samet et al., 2000 

[reanalysis Dominici et al., 2003]; Schwartz, 2005), season-stratified analyses indicated that 

the O3-mortality effect estimates were significant and positive in the warm season, with larger 

effects observed compared to the year-round analyses. The effect estimates from the cool season 

were notably smaller and less significant. In 14 U.S. cities, Schwartz (2005) observed an excess 

mortality risk of 1.0% (95% CI: 0.3, 1.8) per 40 ppb increase in 1-h max O3 during the warm 

season and 0% (95% CI: !1.1, 1.1) during the cold season with a 0-day lag of exposure. Ozone 

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Figure 7-21. All-cause (nonaccidental) O3 excess mortality risk estimates (95% CI) 

by season per standardized increment (see Section 7.1.3.2). Analyses include 

all ages unless otherwise noted. 

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concentrations by season were not presented. Gryparis et al. (2004) estimated an excess risk of 

1.8% (95% CI: 0.99, 3.06) and 0.70% (95% CI: !0.70, 2.17) per 30 ppb increase in 8-h max O3 

at a 0- and 1-day lag during the warm season and cold season, respectively, in 21 European 

cities. The median 8-h max O3 concentrations ranged from 30 to 99 ppb during the warm season 

and 8 to 49 ppb during the cold season. In the U.S. 90 cities study (of which 80 cities had O3 

data available), the winter (December, January, and February) mortality estimate was negative 

(Samet et al., 2000 [reanalysis Dominici et al., 2003]) and was most likely attributable to the 

inverse relationship between O3 and PM in the winter. 

In the U.S. 95 communities study by Bell et al. (2004), no significant difference was found 

between the estimates from all available data and warm-season-only data (for April- October); 

however, no cool-season-only analyses were performed. The warm-season effect estimate using 

the 7-day constrained distributed lag model was 0.78% (95% PI: 0.26, 1.30) excess risk per 

20 ppb increase in 24-h avg O3, compared to 1.04% (95% PI: 0.54, 1.55) calculated using all 

available data. In the 55 communities with year-round O3 data, the all-year effect estimate was 

0.96% (95% PI: 0.32, 1.57). As stated in the previous section, the range of median 

concentrations from the 55 communities monitored all year overlapped considerably with the 

range from the 40 communities with warm-season-only data (see Figure 7-16). 

All three recent meta-analyses (Bell et al., 2005; Ito et al., 2005; Levy et al. 2005), found 

that the warm-season estimates were larger than all-year estimates. In Bell et al. (2005), the 

warm-season estimate was 3.02% (95% PI: 1.45, 4.63), compared to the all-year estimate of 

1.75% (95% PI: 1.10, 2.37). In the subset of 10 cities examined by Ito et al. (2005), the warmseason and all-year estimates were 3.5% (95% CI: 2.1, 4.9) and 2.2% (95% CI: 0.8, 3.6), 

respectively. Similarly, Levy et al. (2005) observed a 3.38% (95% CI: 2.27, 4.42) excess risk in 

the warm season compared to a 1.64% (95% CI: 1.25, 2.03) excess risk using all-year data. 

All results presented are percent excess risk in mortality per standardized increment. 

Studies that conducted analysis by season indicate that O3-mortality risk estimates are 

often larger in the warm season compared to the colder season. The seasonal dependence 

of O3-mortality effects complicates interpretation of O3 risk estimates calculated from year-round 

data without adequate adjustment of temporal trends. 

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7.4.6 Ozone-Mortality Risk Estimates Adjusting for PM Exposure 

The confounding between “winter type” pollution (e.g., CO, SO2, and NO2) and O3 is not 

of great concern, because the peaks of these pollutants do not strongly coincide. The main 

confounders of interest for O3, especially for the northeast United States, are “summer hazetype” pollutants such as acid aerosols and sulfates. Since very few studies included these 

chemical measurements, PM (especially PM2.5) data, may serve as surrogates. However, due to 

the expected high correlation among the constituents of the “summer haze mix,” multipollutant 

models including these pollutants may result in unstable coefficients; and, therefore, 

interpretation of such results requires some caution. 

Figure 7-22 shows the O3 risk estimates with and without adjustment for PM indices using 

all-year data in studies that conducted two-pollutant analyses. Approximately half of the O3 risk 

estimates increased slightly, whereas the other half decreased slightly with the inclusion of PM 

in the models. In general, the O3 -mortality risk estimates were robust to adjustment for PM in 

the models, with the exception of Los Angeles, CA data with PM10 (Kinney et al., 1995) and 

Mexico City data with TSP (Borja-Aburto et al., 1997). 

The U.S. 95 communities study by Bell et al. (2004) examined the sensitivity of acute 

O3-mortality effects to potential confounding by PM10. Restricting analysis to days when 

both O3 and PM10 data were available, the community-specific O3-mortality effect estimates as 

well as the national average results indicated that O3 was robust to adjustment for PM10 (Bell 

et al., 2004). There were insufficient data available to examine potential confounding by PM2.5. 

One study (Lipfert et al., 2000a) reported O3 risk estimates with and without sulfate adjustment. 

Lipfert et al. (2000a) calculated O3 risk estimates based on mean (45 ppb) less background (not 

stated) levels of 1-h max O3 in seven counties in Pennsylvania and New Jersey. The O3 risk 

estimate was not substantially affected by the addition of sulfate in the model (3.2% versus 

3.0% with sulfate) and remained statistically significant. 

Several O3-mortality studies examined the effect of confounding by PM indices in different 

seasons (Figure 7-23). In analyses using all-year data and warm-season only data, O3 risk 

estimates were once again fairly robust to adjustment for PM indices, with values showing both 

slight increases and decreases with the inclusion of PM in the model. In the analyses using coolseason data only, the O3 risk estimates all increased slightly with the adjustment of PM indices, 

although none reached statistical significance. 

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Figure 7-22. All-cause (nonaccidental) O3 excess mortality risk estimates (95% CI) 

with adjustment for PM indices for all-year analyses per standardized 

increment (see Section 7.1.3.2). Analyses include all ages unless 

otherwise noted. 

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Figure 7-23. All-cause (nonaccidental) O3 excess mortality risk estimates (95% CI) 

with adjustment for PM indices by season per standardized increment 

(see Section 7.1.3.2). Analyses include all ages unless otherwise noted. 

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The three recent meta-analyses (Bell et al., 2005; Ito et al., 2005; Levy et al. 2005) all 

examined the influence of PM on O3 risk estimates. No substantial influence was observed in 

any of these studies. In the analysis by Bell et al. (2005), the combined estimate without PM 

adjustment was 1.75% (95% PI: 1.10, 2.37) from 41 estimates, and the combined estimate with 

PM adjustment was 1.95% (95% PI: !0.06, 4.00) from 11 estimates per 20 ppb increase in 24-h 

avg O3. In the meta-analysis of 15 cities by Ito et al. (2005), the combined estimate was 1.6% 

(95% CI: 1.1, 2.2) and 1.5% (95% CI: 0.8, 2.2) per 20 ppb in 24-h avg O3 without and with PM 

adjustment, respectively. The additional time-series analysis of six cities by Ito et al. found that 

the influence of PM by season varied across alternative weather models but was never 

substantial. Levy et al. (2005) examined the regression relationships between O3 and PM 

indices (PM10 and PM2.5) with O3-mortality effect estimates for all year and by season. Positive 

slopes, which might indicate potential confounding, were observed for PM2.5 on O3 risk estimates 

in the summer and all-year periods, but the relationships were weak. The effect of one causal 

variable (i.e., O3) is expected to be overestimated when a second causal variable (e.g., PM) is 

excluded from the analysis, if the two variables are positively correlated and act in the same 

direction. However, the results from these meta-analyses, as well as several single- and 

multiple-city studies, indicate that copollutants generally do not appear to substantially 

confound the association between O3 and mortality. 

7.4.7 Ozone Risk Estimates for Specific Causes of Mortality 

In addition to all-cause mortality, several studies examined broad underlying causes of 

mortality, such as cardiovascular and respiratory causes. The U.S. 95 communities study (1987­

2000) analyzed O3 effect estimates from cardiovascular and respiratory mortality (Bell et al., 

2004). The analysis by Bell et al. (2005) used all available data, which included all-year data 

from 55 communities and warm-season only data from 40 communities. The national average 

estimate from the constrained distributed lag model was slightly greater for cardiopulmonary 

deaths than deaths from all causes, with an excess risk of 1.28% (95% PI: 0.62, 1.97) compared 

to 1.04% (95% PI: 0.54, 1.55) per 20 ppb increase in 24-h avg O3 in the preceding week. 

In a related study, Huang et al. (2005) examined O3 effects on cardiopulmonary mortality 

during the summers (June to September) of 1987 to 1994 in 19 large U.S. cities from the 

NMMAPS database. Figure 7-24 presents the Bayesian city-specific and overall average O3 risk 

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Figure 7-24. City-specific Bayesian estimates and overall average for the percent change 

(95% PI) in cardiopulmonary mortality per 20 ppb increase in 24-h avg O3 in 

the previous week using a constrained distributed lag model for 19 U.S. cities 

(NMMAPS), arranged by size of the effect estimate. Analyses were 

conducted using summer (June to September) data only. 

Source: Derived from Huang et al. (2005). 

estimates for cardiopulmonary mortality per 20 ppb increase in 24-h avg O3 from a constrained 

7-day distributed lag model. The O3 effect estimate was 2.52% (95% PI: 0.94, 4.10) excess risk 

in cardiopulmonary mortality per 20 ppb increase in 24-h avg O3 in the preceding week for the 

combined analysis of all cities. For analyses of summer data, confounding of the O3 effect by 

PM is of concern as daily variations in O3 may be correlated to PM during the summer months. 

Huang et al. observed that when PM10 was included in the model, the O3 effect estimate, on 

average, remained positive and significant. As PM10 measurements were available only every 

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1 to 6 days, only single-day lags were examined. At a 0-day lag, O3 was associated with a 1.47% 

(95% PI: 0.54, 2.39) excess risk versus a 1.49% (95% PI: !0.66, 3.47) excess risk in 

cardiopulmonary mortality in the O3-only model and after adjustment for PM10, respectively. 

The slight sensitivity of the O3 health effects to the inclusion of PM10 in the model may indicate 

a true confounding effect. However, as only the days with PM10 data available were included in 

the analysis, the lack of significance is likely attributable to higher statistical uncertainty due to 

the lack of daily PM10 measurements. 

Figure 7-25 presents effect estimates for associations between O3 and cardiovascular 

mortality for all-year and warm-season analyses. All studies, with the exception of Pönkä et al. 

(1998), showed positive associations between O3 and cardiovascular mortality. The median 

24-h avg O3 was 9 ppb (5th % to 95th %: 2 to 26) in the Helsinki study by Pönkä et al. (1998). 

In addition to Cook County, IL (median 24-h avg O3 of 18 ppb), Moolgavkar (2000) also 

examined the effect of O3 on cardiovascular mortality in Los Angeles County, CA (median 24-h 

avg O3 of 24 ppb) and Maricopa County, AZ (median 24-h avg O3 of 25 ppb). Ozone was not 

found to be associated with cardiovascular or cerebrovascular mortality in these two counties, 

even when data was restricted to the warm season (quantitative risk estimates not presented). 

In a follow-up study, Moolgavkar (2003) analyzed all circulatory system deaths in one broad 

group in order to increase the power of the season-specific analyses using the Cook County and 

Los Angeles County data. In this analysis, O3 concentrations were marginally associated with 

increased circulatory mortality during the summer in Los Angeles County (1.6% [95% CI: !0.2, 

3.5] excess risk per 20 ppb increase in 24-h avg O3). Note that both analyses were performed 

using Poisson GAM with default convergence criteria. 

As with all-cause mortality, there appears to be heterogeneity in the effect estimates across 

studies. The cardiovascular mortality estimate from the meta-analysis by Bell et al. (2005) 

appears to be close to the mode of the effect estimates from the various studies, as shown in 

Figure 7-25. This is expected, given that many of these studies were also included in the 

meta-analysis. Bell et al. (2005) observed that the posterior mean estimate for cardiovascular 

causes (2.23% [95% PI: 1.36, 3.08] excess risk per 20 ppb increase in 24-h avg O3 from 25 

estimates) was slightly larger than that for total mortality (1.75% [95% PI: 1.10, 2.37] excess 

risk from 41 estimates). However, since cardiovascular deaths account for the largest fraction 

(over 40%) of total deaths, it is not surprising that the risk estimates for cardiovascular mortality 

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Figure 7-25. Ozone-associated cardiovascular mortality risk estimates (95% CI) per standardiz

increment (see Section 7.1.3.2). Analyses include all ages unless otherwise noted. 

*Note that Goldberg et al. (2001), Le Tertre et al. (2002a), and Moolgavkar (2001, 2003) performed analyses using Poisson with default convergence criteria. 
are somewhat similar to those from all-cause mortality. Overall, the cardiovascular mortality 

risk estimates in the current literature show consistently positive associations with some 

heterogeneity (most estimates fall within the range of 1 to 8% per 40 ppb increase in 1-h 

avg O3). 

Several studies observed that the risk estimates for the respiratory category were larger 

than the cardiovascular and total nonaccidental categories (e.g., Anderson et al., 1996; Gouveia 

and Fletcher, 2000b; Gryparis et al., 2004; Zmirou et al., 1998). In the European 21 multicities 

study (Gryparis et al., 2004), the warm-season effect estimate for respiratory mortality was 

6.75% (95% CI: 4.38, 9.10) excess risk per 30 ppb increase in 8-h max O3, compared to 2.70% 

(95% CI: 1.29, 4.32) for cardiovascular mortality and 1.82% (95% CI: 0.99, 3.06) for total 

mortality. In contrast, other studies have found that the risk estimates for the respiratory 

category were smaller or even negative, whereas the risk estimates for total or cardiovascular 

categories were positive (e.g., Borja-Aburto et al., 1998; Bremner et al., 1999; Lipfert et al., 

2000a; Morgan et al., 1998b). The apparent inconsistencies across studies may be due in part to 

the differences in model specifications, but they may also reflect the lower statistical power 

associated with the smaller daily counts of the respiratory category (usually accounting for less 

than 10% of total deaths) compared to the larger daily counts for the cardiovascular category 

(approximately 40 to 50% of total deaths). Thus, an examination of the differences in risk 

estimates across specific causes requires a large population and/or a long period of data 

collection. In the meta-analysis by Bell et al. (2005), which combined 23 estimates from 

17 studies for respiratory mortality, the effect estimate for respiratory causes was smaller (0.94% 

[95% PI: !1.02, 2.96] excess risk per 20 ppb increase in 24-h avg O3) compared to the estimates 

for total mortality (1.75% excess risk) and cardiovascular mortality (2.23% excess risk). 

The analyses of a 9-year data set for the whole population of the Netherlands 

(population = 14.8 million) provided risk estimates for more specific causes of mortality, 

including COPD, pneumonia, and subcategories of cardiovascular causes (Hoek et al., 2000, 

2001; reanalysis Hoek, 2003). The median 8-h avg O3 (12 p.m.-8 p.m.) was 24 ppb (range 1 to 

117). The effect estimate for total nonaccidental mortality was 1.6% (95% CI: 0.9, 2.4) excess 

risk per 30 ppb increase in 8-h avg O3. In comparison, the excess risk estimates for pneumonia 

and COPD were 5.6% (95% CI: 1.8, 9.5) and 0.8% (95% CI: !2.4, 4.2), respectively. The 

effect estimates for some of the cardiovascular subcategories, including heart failure (3.8% [95% 

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CI: 0.5, 7.3]) and thrombosis-related disease (6.0% [95% CI: 1.1, 10.8]), showed greater risk 

estimates than that for total mortality. However, these associations were not specific to O3. 

! For example, most of the pollutants examined, including PM10, BS, SO2, NO2, CO and NO3 , 

were associated with pneumonia. Therefore, it is difficult to make a causal inference specific 

to O3 on the basis of these results. 

De Leon et al. (2003) examined the role of contributing respiratory causes in the 

association between air pollution and nonrespiratory mortality (circulatory and cancer) in 

New York City during the period of 1985 to 1994. The mean 24-h avg O3 concentration was 

21.6 ppb (5th % to 95th %: 7.0 to 45.0). The main finding of this study was that the estimated 

excess mortality risks for PM10 were higher for nonrespiratory deaths that had contributing 

respiratory causes compared to deaths without contributing respiratory causes in the older age 

(75+ years) group. This pattern was also seen for CO and SO2, but not for O3. Therefore, this 

study did not suggest a role of contributing respiratory causes in the association between O3 and 

nonrespiratory causes of deaths. 

In summary, several single-city studies observed positive associations between ambient O3 

concentrations and cardiovascular mortality. In addition, a meta-analysis that examined specific 

causes of mortality found that the cardiovascular mortality risk estimates were higher than those 

for total mortality. The findings regarding the effect size for respiratory mortality have been less 

consistent, possibly because of lower statistical power in this subcategory of mortality. 

7.4.8 Ozone-Mortality Risk Estimates for Specific Subpopulations 

Some studies have examined O3-mortality risk estimates in potentially susceptible 

subpopulations, such as those persons with underlying cardiopulmonary disease. Sunyer et al. 

(2002) examined the association between air pollution and mortality in a cohort of patients (467 

men and 611 women) with severe asthma in Barcelona, Spain during the period of 1986 to 1995. 

A case-crossover study design was used to estimate excess odds of mortality adjusting for 

weather and epidemics in three groups: (1) those who had only one asthma emergency 

department visit; (2) those who had more than one asthma emergency department visit; and 

(3) those who had more than one asthma and COPD emergency department visit. Those with 

more than one asthma emergency department visit showed the strongest associations with the 

examined air pollutants, with NO2 being the most significant predictor, followed by O3. Sunyer 

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et al. (2002) reported a significant association between O3 and all-cause deaths for this group 

during the warm season, with an odds ratio of 2.82 (95% CI: 1.15, 6.87) per 40 ppb increase in 

1-h max O3, compared to an odds ratio of 1.03 (95% CI: 0.60, 1.78) for those with only one 

asthma emergency department visit and 1.08 (95% CI: 0.60, 1.92) for the group with a 

concurrent diagnosis of COPD. The median 1-h max O3 concentration for all-year data was 

36 ppb, but seasonal O3 concentrations were not presented. In another Barcelona study, Saez 

et al. (1999) examined asthma mortality death among persons aged 2 to 45 years. Once again, 

O3 and NO2 were the only air pollutants found to be significantly associated with asthma 

mortality death. While the similarity of the patterns of associations between O3 and NO2 makes 

it difficult to speculate on the specific causal role of O3, the results of these studies suggest that 

individuals with severe asthma may make up a subpopulation that is sensitive to these pollutants. 

Sunyer and Basagna (2001) also performed an analysis of emergency department visits for 

individuals with COPD. The results from this study suggested that PM10, but not gases were 

associated with mortality risks for the COPD cohort. The mean O3 concentration was not 

provided (IQR was 11 ppb for 1-h max O3). However, a Mexico City study by Téllez-Rojo et al. 

(2000) observed a significant association between COPD mortality and O3, as well as PM10, 

among patients living outside a medical unit. For a cumulative 5-day lag, an excess risk of 

15.6% (95% CI: 4.0, 28.4) per 40 ppb increase in 1-h max O3 was observed for COPD mortality. 

The mean 1-h max O3 was 69.4 ppb (SD 17.2). 

Goldberg et al. (2003) investigated associations between air pollution and daily mortality 

with congestive heart failure as the underlying cause of death in patients aged 65 years or more 

in Montreal, Quebec during the period of 1984 to 1993. The mean 24-h avg O3 was 15.0 ppb 

(SD 8.8). Analysis was stratified into two groups, those whose underlying cause of death was 

congestive heart failure and those with a diagnosis of congestive heart failure one year before 

their death. No association was found between daily mortality for congestive heart failure and 

any pollutants. However, associations were found between daily mortality and coefficient of 

haze, SO2, and NO2 among patients who were classified as having congestive heart failure before 

death. In the case of O3, positive risk estimates were observed for year-round and warm-season 

data; however, the results were not statistically significant. Although the 10-year study period 

for this data was long, the daily mean death counts for the specific subcategory chosen was 

relatively small (0.7/day for mortality with congestive heart failure as underlying cause of death 

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and 4.0/day for total mortality in patients previously diagnosed with congestive heart failure), 

thus limiting the power of the study. 

In the meta-analysis by Bell et al. (2005), a combined estimate was obtained for the elderly 

population (age 64 years and older or 65 years and older) using 10 estimates from 9 studies. The 

posterior mean estimate for the elderly category (2.92% [95% PI: 1.34, 4.51] per 20 ppb 

increase in 24-h avg O3) was larger than that from all ages (1.75% [95% PI: 1.10, 2.37] from 

41 estimates). The results from the Bell et al. (2005) meta-analysis suggest that the elderly 

population may be particularly susceptible to O3-related mortality. 

Few studies have examined O3-mortality effects for specific subpopulations. Among those 

that investigated the effect of air pollution in populations with underlying cardiopulmonary 

diseases, associations were not unique to O3 but, rather, were shared with other pollutants. There 

is suggestive evidence that severe asthmatics may be susceptible to the mortality effects 

associated with NO2 and O3. In addition, the meta-analysis by Bell et al. (2005) suggests that the 

elderly population may be more affected by O3. 

7.4.9 Summary of Acute Ozone Effects on Mortality 

C A substantial body of new data on acute mortality effects of O3 has emerged since the 

1996 O3 AQCD. While uncertainties remain in some areas, it can be concluded that 

robust associations have been identified between various measures of daily O3 

concentrations and increased risk of mortality. Most of the single-pollutant model 

estimates from single-city studies fall in the range between 0.5 to 5% excess deaths per 

standardized increment. The corresponding summary estimates in large U.S. multicity 

studies ranged between 0.5 to 1%, with some studies noting heterogeneity across cities 

and studies. These associations could not be readily explained by confounding due to 

time, weather, nor copollutants. Differences between populations as well as model 

specifications likely contributed to some of the observed heterogeneity in risk 

estimates across studies. 

C Most studies calculated O3-mortality risk estimates using all-year data. The results 

from studies that conducted stratified analyses by season suggested that the O3 risk 

estimates were larger in the warm season. Some of the risk estimates in the cool 

season were negative, possibly reflecting the negative correlation between low-level 

O3 and PM (and/or other primary pollutants) during that season. Thus, even with 

adjustment for temporal trends, the O3 risk estimates obtained for year-round data may 

be misleading. In locations with considerable seasonal variation, season-specific 

analyses may better elucidate the effect of O3 on mortality. 

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C The majority of the available O3-mortality risk estimates were computed for a singleday lag. Choosing the optimal lag out of several lags examined may bias the singleday risk estimate upward. However, recent findings from the largest U.S. 95 

communities study indicated that a strong association between O3 and mortality was 

observed with a 7-day distributed lag model. Thus, it is possible that the effect of 

acute O3 exposure on mortality persists over several days. Since the parameters 

estimated from single-day lag versus multiday lag models are not the same, 

comparison of these results are difficult. 

C Some studies examined specific subcategories of mortality, but most of these studies 

had limited statistical power to detect associations due to the small daily mortality 

counts. A recent meta-analysis indicated that there was a slightly greater risk of 

cardiovascular mortality compared to total mortality. 

C Few studies examined the effect of O3 on mortality in subpopulations with underlying 

cardiopulmonary diseases. Similar to cause-specific mortality, these populationspecific studies had limited statistical power to detect associations. 

7.5 EFFECTS OF CHRONIC OZONE EXPOSURE 

7.5.1 Summary of Key Findings on Studies of Health Effects and Chronic 

Ozone Exposure from the 1996 Ozone AQCD 

The 1996 O3 AQCD concluded that there was insufficient evidence from the limited 

number of studies to determine whether long-term ambient O3 exposures resulted in chronic 

health effects. However, the aggregate evidence suggested that chronic O3 exposure, along with 

other environmental factors, could be responsible for health effects in exposed populations. 

7.5.2 Introduction to Morbidity Effects of Chronic Ozone Exposure 

Several new longitudinal epidemiologic investigations have yielded information on health 

effects of long-term O3 exposures. Epidemiologic interest in investigating long-term effects has 

been motivated by several considerations. Animal toxicology studies carried out from the late 

1980s onward demonstrated that long-term exposures can result in permanent changes in the 

small airways of the lung, including remodeling of the airway architecture (specifically the distal 

airways and centriacinar region) and deposition of collagen, as discussed earlier in Chapter 5. 

These changes result from the damage and repair processes that occur with repeated exposure. 

Indices of fibrosis also were found to persist after exposure in some of the studies. Collectively, 

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these findings provide a potential pathophysiologic basis for the changes in airway function 

observed in children in longitudinal studies. Seasonal ambient patterns of exposure may be of 

greater concern than continuous daily exposure. In the classical study by Tyler et al. (1988), 

young monkeys with seasonal exposure to O3, but not those with daily exposure, experienced 

increases in total lung collagen content, chest wall compliance, and inspiratory capacity, 

suggesting a delay in lung maturation in seasonally-exposed animals. 

Controlled human exposure studies clearly demonstrated acute inflammation in the lung at 

ambient exposure levels. Epidemiologic studies could examine whether repeated exposures over 

multiple episode periods and/or multiple years would lead to persistent inflammation and result 

in damage to the human lung, especially in the small, terminal bronchiolar regions where 

vulnerability is greatest. However, the challenges to addressing these issues in epidemiologic 

studies are formidable, and as a result there exists relatively limited literature in this area. Longterm O3 concentrations tend to be correlated with long-term concentrations of other pollutants, 

making specific attribution difficult. Subtle pulmonary effects require health outcome measures 

that are sensitive, and must usually be directly collected from individual human subjects, rather 

than from administrative data bases. Although these factors make chronic studies difficult and 

expensive to conduct, efforts must be made to design studies with adequate power to examine 

the hypothesis being tested. Epidemiologic studies have the potential to provide important new 

insights on the links between chronic exposure to O3 and the occurrence of human health effects. 

This section reviews studies published from 1996 onward in which health effects were 

tested in relation to extended O3 exposures ranging from several weeks to many years 

(Table AX7-6 in the Annex 7, Section AX7.1). The available literature falls into four general 

categories: (1) studies examining seasonal changes in lung function as related to O3 exposures 

in peak season; (2) studies addressing smaller increases in lung function during childhood or 

decline of lung function beyond childhood in relation to long-term O3 exposures; (3) studies 

addressing respiratory inflammation in high versus low exposure groups or time periods; and 

(4) studies addressing longitudinal and cross-sectional associations between long-term O3 

exposures and asthma development and prevalence. 

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7.5.3 Seasonal Ozone Effects on Lung Function 

While it has been well-documented in both chamber and field studies that daily, multihour 

exposures to O3 result in transient declines in lung function, much less is known about the effects 

of repeated exposures to O3 over extended periods on lung function. Several new studies 

reported during the past decade have examined lung function changes over seasonal time periods 

with differing levels of O3 exposures (Frischer et al., 1999; Horak et al., 2002a,b; Ihorst et al., 

2004; Kinney and Lippmann, 2000; Kopp et al., 2000). The seasonal effects of O3 are examined 

first in this section. The next section then discusses effects of O3 exposures over years, as 

opposed to over seasons, in addition to multiyear analyses of seasonal studies. 

In a large Austrian study, Frischer et al. (1999) carried out repeated lung function 

measurements in 1,150 school children (mean age 7.8 years) from nine towns that differed in 

mean O3 levels. Lung function was measured in the spring and fall over a 3-year period from 

1994 to 1996, yielding six measurements per child. Mean summertime 24-h avg O3 

concentrations ranged from 32.4 to 37.3 ppb during the three summers. Growth-related 

increases in lung function over the summer season were reduced in relation to seasonal mean O3 

levels. Ozone was associated with a change of !156.6 mL (95% CI: !209.5, !103.7) (central 

estimate: !0.029 mL/day/ppb × 90 days/year × 3 years × 20 ppb) in FEV1 increase for each 

20 ppb increase in mean 24-h avg O3 concentrations over the three summers and !129.6 mL 

(95% CI: !193.1, !66.1) over the three winters. When analyses were restricted to children who 

had spent the whole summer period in their community, the changes were greater, with an O3­

related !183.6 mL (95% CI: !278.9, !88.3) change in FEV1 increase over three summers. 

Other pollutants (PM10, SO2, and NO2) had less consistent associations with changes in lung 

function. Horak et al. (2002a,b) extended the study of Frischer et al. (1999) with an additional 

year of data and found that seasonal mean O3 was associated with a negative effect on increases 

in lung function, confirming results from the previous 3-year study. In an editorial, Tager (1999) 

stated that the Frischer et al. (1999) data provided the first prospective evidence of an association 

between exposure to ambient air pollution and alterations in lung function in children. Tager 

further noted that the prospective study design represented a substantial improvement over data 

derived from cross-sectional studies and should be emulated. However, Tager also cautioned 

that it was difficult to attribute the reported effects to O3 alone independently of copollutants. 

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Kopp et al. (2000) reported smaller increases in lung function among a cohort of 797 

children from 10 communities in Austria and southwestern Germany who were exposed to high 

levels of ambient O3 (mean O3 concentration of 44 to 52 ppb from April to October 1994). 

Children residing in lower O3 (24 to 33 ppb) areas experienced a 43 mL increase in FEV1, 

whereas those in high O3 areas only manifested a 16 mL increase during the summer of 1994. 

Similar results were found in data for the summer of 1995. In another study of 15 communities 

in Austria and southwestern Germany, Ihorst et al. (2004) examined 2,153 children with a 

median age of 7.6 years and reported summer pulmonary function results which indicated that 

significantly lower FVC and FEV1 increases were associated with higher O3 exposures in the 

summer, but not in the winter. Semi-annual mean O3 concentrations ranged from 22 to 54 ppb 

during the summer and 4 to 36 ppb during the winter. 

In a pilot study (Kinney and Lippmann, 2000), 72 nonsmoking adults (mean age 20 years) 

from the second year class of students at the U.S. Military Academy in West Point, NY provided 

two lung function measurements, one before and one after a 5-week long summer training 

program at four locations. There was a greater decline in FEV1 among students at the Fort Dix 

location (78 mL) as compared to students at the other locations (31 mL). Ozone levels at Fort 

Dix averaged 71 ppb (mean of daily 1-h max O3) over the summer training period versus mean 

values of 55 to 62 ppb at the other three locations. In addition to the higher mean O3 level, Fort 

Dix had greater peak O3 values (23 hours >100 ppb) compared to the other locations (1 hour 

>100 ppb). Ambient levels of other pollutants, PM10 and SO2, were relatively low during the 

study and did not vary across the four sites. Although the conclusions that can be drawn are 

limited by the small study size, the results appear to be consistent with a seasonal decline in lung 

function that may be due, in part, to O3 exposures. An exploratory observation from this study 

was that there appeared to be a larger decline for those subjects who completed their postsummer lung function measurements within two weeks after returning from training compared to 

those measured 3 to 4 weeks after training, results consistent with some degree of rebound of 

function following the summer exposure period. 

Collectively, the above studies indicate that seasonal O3 exposure is associated with 

smaller increases in lung function in children. The study by Kinney and Lippman (2000) also 

provides some limited evidence suggesting that seasonal O3 also may affect lung function in 

young adults. 

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7.5.4 Chronic Ozone Exposure Effects on Lung Function and 

Respiratory Symptoms 

Lung capacity grows during childhood and adolescence as body size increases, reaches 

a maximum during the 20s, and then begins to decline steadily and progressively with age. 

There has long been concern that long-term exposure to air pollution might lead to slower 

growth in lung capacity, diminished maximally attained capacity, and/or more rapid decline 

in capacity with age. The concern arises by analogy with cigarette smoking, where it is welldocumented that lung function declines more rapidly with age in a dose-dependent manner 

among adults who smoke cigarettes. Adults who stop smoking return to a normal rate of decline 

in capacity, although there is no evidence that they regain the capacity previously lost due to 

smoking (Burchfiel et al., 1995). Because O3 is a strong respiratory irritant and is associated 

with acute lung function declines as well as inflammation and re-structuring of the respiratory 

airways, it seems plausible that there might be a negative impact of long-term O3 exposures on 

lung function. Exposures that negatively affect increases in lung function during childhood, in 

particular, might have greater long-term risks. Thus, studies of effects on the rates of increases 

in lung function in children are especially important. 

Several studies published over the past decade have examined the relationship between 

long-term O3 exposure and lung function. The most extensive and robust study of respiratory 

effects in relation to long-term air pollution exposures among children in the United States is the 

Children’s Health Study carried out in 12 communities of southern California starting in 1993 

(Avol et al., 2001; Gauderman et al., 2000, 2002, 2004a,b; Peters et al., 1999a,b). The first 

cohort included children from the fourth, seventh, and tenth grades. A total of 3,676 students 

completed questionnaires regarding their lifetime residential histories, historic and current health 

status, residential characteristics, and physical activity. Among those students, 3,293 also 

performed pulmonary function tests at the time of enrollment. Peters et al. (1999a) examined the 

relationship between long-term (1986 to 1990) O3 exposures and self-reports of respiratory 

symptoms and asthma in a cross-sectional analysis. Mean 1-h max O3 levels from 1986 to 1990 

ranged from 30.2 ppb to 109.2 ppb across the 12 communities. For outcomes of current asthma, 

bronchitis, cough, and wheeze, the reported odds ratios were 0.95 (95% CI: 0.70,1.29), 1.14 

(95% CI: 0.84, 1.55), 0.98 (95% CI: 0.82, 1.17), and 1.08 (95% CI: 0.87, 1.35), respectively, 

per 40 ppb increase in 1-h max O3. In another cross-sectional analysis of the relationship 

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FEF

between lung function at baseline and levels of air pollution in the community, there was 

evidence that annual mean O3 levels were associated with decreased FVC, FEV1, PEF, and 

25-75 (the latter two being statistically significant) among females but not males (Peters et al., 

1999b). 

Avol et al. (2001) examined 110 children from the first cohort who had moved from the 

participating communities in southern California to other communities to determine whether 

changes in air quality caused by relocation were associated with changes in annual increases in 

lung function. Negative effects of O3 were observed for all spirometric parameters except FEV1, 

but the associations were not as strong as those observed for PM10. 

To examine the association between long-term exposure to air pollution and changes in 

lung function, fourth graders (n = 1,759) from the first cohort were followed over eight years 

(Gauderman et al., 2000, 2004a,b). During the follow-up period from 1993 to 2001, there was 

an attrition of approximately 10% of subjects per year, resulting in 747 subjects being tested in 

2001. The longitudinal pulmonary function data for each child was related to average air 

pollution levels in each study community using a multistage regression approach. During the 

monitoring period, there was substantial variation in the average levels of various air pollutants 

across the 12 communities, but relatively little year-to-year variation within each community. 

However, O3 levels did not vary widely across the communities in comparison to the other 

pollutants. The mean annual average of 8-h avg O3 concentrations ranged from approximately 

30 ppb in Long Beach to 65 ppb in Lake Arrowhead from 1994 to 2000. Average levels of O3 

were not significantly correlated across communities with any other study pollutant. Analyses 

indicated that there was no evidence that either 8-h avg O3 (10 a.m. to 6 p.m.) or 24-h avg O3 

was associated with any measure of lung function growth over a 4-year (age 10 to 14 years; 

Gauderman et al., 2000) or 8-year (age 10 to 18 years; Gauderman et al., 2004) period. 

However, most of the other pollutants examined (including PM2.5, NO2, acid vapor, and 

elemental carbon) were found to be significantly associated with reduced growth in lung 

function. Also, a clinically low FEV1 (i.e., FEV1 <80% of predicted value) at age 18 years was 

reported to be correlated with all pollutants examined except for O3 (Figure 7-26). 

The Children’s Health Study enrolled a second cohort of fourth graders (n = 1,678) in 1996 

(Gauderman et al., 2002). While the strongest associations with negative lung function growth 

were observed with acid vapors in this cohort, children from communities with higher 4-year 

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Figure 7-26. Proportion of 18-year-olds with a FEV1 below 80% of the predicted value 

plotted against the average levels of pollutants from 1994 through 2000 in 

the 12 southern California communities of the Children’s Health Study. 

AL = Alpine; AT = Atascadero; LA = Lake Arrowhead; LB = Long Beach; LE = Lake Elsinore;

LM = Lompoc; LN = Lancaster; ML = Mira Loma; RV = Riverside; SD = San Dimas;

SM = Santa Maria; UP = Upland

Source: Derived from Gauderman et al. (2004). 

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average O3 levels also experienced smaller increases in various lung function parameters. The 

strongest effect of O3 was on PEF, with children from the least-polluted community having a 

1.21% (95% CI: 0.36, 2.06) greater increase in PEF compared to those from the most-polluted 

communities. In two-pollutant models, only 8-h avg O3 and NO2 were significant joint 

predictors of FEV1 and maximal midexpiratory flow (MMEF). The range of mean annual 

average 8-h avg O3 levels across the 12 communities was similar to that for the first cohort, 27 to 

67 ppb. 

In both cohorts of fourth graders, stratified analyses by time spent outdoors indicated a 

stronger association between long-term O3 exposure and smaller increases in lung function 

during the 4-year follow-up in children who spent more time outdoors, as shown in Table 7-3. 

Jedrychowski et al. (2001) had reported a link between repeated respiratory symptoms and 

smaller lung function increases. Gauderman et al. (2002), therefore, suggested that the 

observation of reduced increases in lung function with increasing annual average air pollution 

might be a consequence of repeated acute respiratory events after short-term increases in 

pollutant levels. The findings of larger deficits in children who spend more time outdoors in the 

afternoon adds some support for this possibility. 

Although results from the second cohort of children are supportive of an association, the 

definitive 8-year follow-up analysis of the first cohort (Gauderman et al., 2004a) provides little 

evidence that long-term exposure to ambient O3 at current levels is associated with significant 

deficits in the growth rate of lung function in children. It should be noted, however, that O3 

exposure prior to the study period is unknown. In addition, there was only about a 2- to 2.5-fold 

difference in O3 concentrations from the least to most polluted communities (mean annual 

average of 8-h avg O3 ranged from 30 to 65 ppb), versus the ranges observed for the other 

pollutants (which had 4- to 8-fold differences in concentrations). Finally, the results from the 

stratified analyses by time spent outdoors indicate the potential for substantial misclassification 

of long-term O3 exposure. The Children’s Health Study, like most long-term epidemiologic 

studies, estimated O3 exposure using centrally-located ambient monitors. Gonzales et al. (2003) 

and Künzli et al. (1997) evaluated the use of retrospective questionnaires to reconstruct past 

time-activity and location pattern information. Both studies found that questionnaires or activity 

diaries might improve the assessment of chronic exposure in epidemiologic studies. 

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Table 7-3. Difference in Annual Percent Increases in Lung Function from the Least to 

the Most Polluted Community in the Children’s Health Study by Time Spent Outdoors a 

Lung Function 

Parameter Cohort b 

More Time Outdoors c 

% Change (95% CI) d 

Less Time Outdoors c 

% Change (95% CI) d

 FVC Cohort 1 !0.02% (!0.57, 0.54) !0.04% (!0.45, 0.37) 

Cohort 2 !0.57% (!1.03, !0.09) !0.06% (!0.76, 0.66)

 FEV1 Cohort 1 !0.25% (!1.18, 0.68) !0.05% (!0.58, 0.49) 

Cohort 2 !0.68% (!1.36, 0.00) !0.29% (!1.02, 0.46)

 MMEF Cohort 1 !0.55% (!2.08, 1.01) 0.23% (0.89, 1.36) 

Cohort 2 !0.48% (!1.71, 0.78) !0.80% (!2.07, 0.50)

 PEF Cohort 1 !0.77% (!2.03, 0.52) 0.25% (!0.65, 1.16) 

Cohort 2 !1.33% (!2.43, !0.24) !0.71% (!1.71, 0.30) 

a Results are derived from Gauderman et al. (2002).

b Cohort 1 includes children enrolled in 1993 as 4th graders and followed through 1997 (n = 1,457). 

Cohort 2 includes children enrolled in 1996 as 4th graders and followed through 2000 (n = 1,678).

c More or less time outdoors is based on reported time spent outdoors during weekday afternoons. 

 Subjects were split into the two groups on the basis of the median reported time outdoors within each cohort. d Percent change in lung function is per 30 ppb increase in 8-h avg O3 (10 a.m.-6 p.m.).

In a study of 15 communities in Austria and Germany, Ihorst et al. (2004) found no 

associations between increases in lung function and mean summer O3 levels for FVC and FEV1 

over a 3.5-year period, in contrast to the significant seasonal effects discussed in the earlier 

section above. Unlike the reduced increases in lung function parameters over the first two 

summers among children in high O3 areas noted above, a greater increase was observed by Ihorst 

et al. during the third summer and no difference was observed during the fourth summer. The 

mean O3 concentration during the first two summers across the 15 communities was similar to 

that during the last two summers, 36.6 ppb and 35.1 ppb, respectively. The authors concluded 

that medium-term effects on school children lung function were possibly present but were not 

detected over a 3- to 5-year period due to partial reversibility. A study by Frischer et al. (1999) 

showed results similar to the Ihorst et al. (2004) study. Although O3 was related to smaller 

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increases in lung function when three years of data were analyzed collectively, the magnitude 

and direction of the effect changed throughout the years. Ozone was associated with a change of 

!34.0 mL (95% CI: !58.7, !9.3) in FEV1 increase in the first year, compared to +7.3 mL (95% 

CI: !20.8, 35.6) in the third year for each 20 ppb increase in mean 24-h avg O3. Mean 

summertime O3 concentrations decreased slightly over the three years; the mean levels being 

37.3 ppb, 35.4 ppb, and 32.4 ppb during the first, second, and third years, respectively.

A study by Gong et al. (1998b) examined lung function changes in 164 nonsmoking adults 

(mean age 45 years) from a high O3 community in southern California, recruited from a cohort 

of 208 who had been tested on two previous occasions. In the earlier analysis by Detels et al. 

(1987), a significant decline in lung function was observed from 1977/1978 to 1982/1983. 

In contrast, Gong et al. (1998b) observed a slight increase in FVC and FEV1 from 1982/1983 to 

1986/1987. The mean annual average 1-h max O3 level was 110 ppb from 1972 to 1982. For 

1983 to 1989, the annual average 1-h max O3 levels ranged from 103 to 134 ppb. A consistent 

decline in FEV1/FVC ratio was observed at all three time points (p < 0.0001). Among the 

45 subjects who further participated in the controlled exposure study (0.40 ppm O3 over 2 hours 

with intermittent exercise), declines in FEV1 and FVC were associated with acute O3 exposure. 

However, the acute changes in lung function were not associated with long-term changes in lung 

function over a decade. 

Evidence for a relationship between long-term O3 exposures and decrements in maximally 

attained lung function was observed in a nationwide cohort of 520 first year students at Yale 

College in New Haven, CT (Galizia and Kinney 1999; Kinney et al., 1998). Each student 

performed one lung function test in the spring of their first year at college. Ozone exposures 

were estimated by linking 10-year average summer season (June to August) 1-h max O3 levels at 

the nearest monitoring station to the residential locations reported each year from birth to the 

time of measurement. The mean 10-year average 1-h max O3 was 61.2 ppb (range 13-185). 

Students who had lived four or more years in areas with long-term mean O3 levels above 80 ppb 

had significantly lower FEV1 (!3.07% [95% CI: !0.22, !5.92]) and FEF25-75 (!8.11% [95% CI: 

!2.32, !13.90]) compared to their classmates with lower long-term O3 exposures. Stratification 

by gender indicated much larger effect estimates for males than for females, which might reflect 

higher outdoor activity levels and corresponding higher O3 exposure during childhood. Ozone 

was the only air pollutant examined in this study. The authors noted, therefore, that it was not 

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possible to determine whether, and to what extent, the observed effects are due to O3 alone or O3 

in combination with unmeasured copollutants. 

FEF

A similar study of 130 first year college freshmen at the University of California at 

Berkeley also reported significant effects of O3 on lung function (Künzli et al., 1997; Tager 

et al., 1998). Enrollment was limited to students from either the San Francisco or Los Angeles, 

CA metropolitan areas. Lifetime monthly average 8-h avg O3 (10 a.m.-6 p.m.) concentrations 

ranged from 16 to 33 ppb for San Francisco residents and 25 to 74 ppb for Los Angeles 

residents. After controlling for city of origin, long-term O3 exposures were found to be 

associated with declines in FEF25-75 and FEF75 (forced expiratory flow after 75% of FVC has 

been exhaled). No effects were seen for PM10 and NO2. Künzli et al. (1997) noted that 

significant changes in these mid- and end-expiratory flow measures could be considered early 

indicators for pathologic changes that might ultimately progress to COPD, as evidenced by 

animal studies that showed that the primary site of O3 injury in the lung was the centriacinar 

region (see Chapter 5). This study was repeated in 2000 to 2002, with 255 freshmen who were 

life-long residents of the Los Angeles or San Francisco areas (Tager et al., 2005). Lifetime 

monthly average 8-h avg O3 levels ranged from 18 to 65 ppb for the entire cohort, with medians 

of 36 ppb for the men and 33 ppb for the women. In contrast to results from the first cohort, 

associations between long-term O3 exposure and declines in FEF25-75 and FEF75 were not 

observed in the second cohort of freshmen. However, when the analysis was stratified by gender 

and an interaction term for intrinsic airway size (FEF25-75/FVC ratio) was included in the model, 

lifetime exposure to O3, as well as PM10 and NO2, was found to be associated with decreased 

25-75 and FEF75 for both men and women. The adverse impact of O3 exposure decreased with 

increasing intrinsic airway size. Also, in multipollutant models including all three pollutants, 

meaningful changes were not observed in the O3 effect estimates, but the PM10 and NO2 

estimates were reduced substantially. 

The cross-sectional studies in college freshmen provide suggestive evidence that young 

adults who have grown up in high O3 communities may have reduced lung function compared to 

those from low O3 communities. However, attributing the effects specifically to O3 is difficult, 

given that the effects of coexposures to other ambient air pollutants have not necessarily been 

adequately addressed in these cross-sectional long-term exposure studies. Thus, results of the 

longitudinal southern California Children’s Health Study, as well as those from the European 

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studies, provide little evidence for impacts of long-term, relatively low-level O3 exposures on 

lung function development in children. On the other hand, these studies appear to indicate likely 

strong effects on lung function growth for long-term exposures to PM2.5 and/or certain other 

copollutants. 

7.5.5 Chronic Ozone Exposure and Respiratory Inflammation 

As noted in Chapter 6, controlled human exposure chamber studies have demonstrated that 

brief (2 to 6.6 hours) exposures to O3 while exercising result in inflammation in the lung, 

including the alveolar region where gas exchange takes place. This acute exposure effect is 

potentially important for effects of chronic exposure, because repeated inflammation can result 

in the release of substances from inflammatory cells that can damage the sensitive cells lining 

the lung. Over extended periods, repeated insults of this kind can lead to permanent damage to 

and restructuring of the small airways and alveoli. In addition, since inflammation is a 

fundamental feature of asthma, there is concern that O3-induced inflammation can exacerbate 

existing asthma or perhaps promote the development of asthma among genetically pre-disposed 

individuals. Several studies are discussed next, examining different outcomes related to 

inflammation. 

In a study by Kinney et al. (1996b), bronchoalveolar lavage fluids were collected in the 

summer and winter from a group of 19 adult joggers living and working on an island in 

New York City harbor. The mean 1-h max O3 concentrations for a 3-month period were 58 ppb 

(maximum 110) in the summer and 32 ppb (maximum 64) in the winter. PM10 and NO2 

concentrations were similar across the two seasons. There was little evidence for acute 

inflammation in bronchoalveolar lavage fluids collected during the summer as compared to that 

collected from the same subjects in the winter. However, there was evidence of enhanced cell 

damage, as measured by lactate dehydrogenase, in the summer lavage fluids. These results 

indicate that acute inflammatory responses may diminish with repeated exposures over the 

course of a summer (as also demonstrated by multiday chamber exposures; see Chapter 6, 

Section 6.9), but cell damage may be ongoing. 

In a cross-sectional study by Frischer et al. (2001), urinary eosinophil protein was analyzed 

as a marker of eosinophil activation in 877 school children living in nine Austrian communities 

with varying ranges of O3 exposure levels. The results indicated that O3 exposure was 

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significantly associated with eosinophil inflammation after adjusting for gender, site, and atopy. 

The mean 30-day average O3 concentration before sample collection was 31.6 ppb (5th % to 

95th%: 11.8 to 51.5). 

Pollution effects in the nose can be viewed as a potential surrogate measure for effects that 

may occur in the lungs, though doses to nasal tissues are usually higher for a given pollutant 

concentration. As discussed in Chapter 5, morphological effects of O3 on the upper respiratory 

tract can indicate quantitative changes in the nasal transitional respiratory epithelium. The 

persistent nature of the O3-induced mucous cell metaplasia in rats, as discussed in Chapter 5, 

suggests that O3 exposure may have the potential to induce similar long-lasting alterations in the 

airways of humans. In a cross-sectional cohort study by Calderón-Garcidueñas et al. (1995), 

nasal lavage samples collected from children living in Mexico City (n = 38) were compared to 

those from children living in a clean coastal town (n = 28). In Mexico City, the 1-h avg O3 

concentrations exceeded 120 ppb for 4.4 h/day, while O3 levels were not detectable in the coastal 

town. Mexico City children were examined four times within a 1-month period. Nasal 

cytologies revealed that all Mexico City children had abnormal nasal mucosae, including 

mucosal atrophy, marked decreases in the numbers of ciliated-type cells and goblet cells, 

and squamous metaplasia. Significantly higher nasal polymorphonuclear leukocyte counts 

(p < 0.001) and expression of human complement receptor type 3 (CD11b) (p < 0.001) were also 

observed in Mexico City children compared to the control children. Analyses using repeated 

measurements indicated that the inflammatory response did not appear to correlate with the 

previous day’s O3 concentration, suggesting that there might be a competing inflammatory 

mechanism at the bronchoalveolar level with structural injury following acute exposure. 

Limited epidemiologic research has been conducted relating long-term O3 exposure to 

inflammation. In the Mexico City study (Calderon-Garciduenas et al., 1995), specific attribution 

of the nasal abnormalities to long-term O3 exposure is difficult, given the complex pollutant 

mixture present in the ambient air. However, the inflammatory changes such as increased 

eosinophil levels observed in the Austrian study (Frischer et al., 2001) would be consistent with 

known effects of O3. 

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7.5.6 Risk of Asthma Development 

SO

Recent longitudinal cohort studies have reported associations between long-term O3 

exposures and the onset of asthma (McConnell et al., 2002; McDonnell et al., 1999). Significant 

associations between long-term O3 exposure and new cases of asthma among adult males were 

observed in a cohort of nonsmoking adults in California (Greer et al., 1993; McDonnell et al., 

1999). The Adventist Health and Smog (AHSMOG) study cohort of 3,914 (age 27 to 87 years, 

36% male) was drawn from nonsmoking, non-Hispanic white California Seventh Day 

Adventists, who were surveyed in 1977, 1987, and 1992. To be eligible, subjects had to have 

lived 10 or more years within 5 miles of their current residence in 1977. Residences from 1977 

onward were followed and linked in time and space to interpolate concentrations of O3, PM10, 

4

2!

, SO2, and NO2. New asthma cases were defined as self-reported doctor-diagnosed asthma 

at either the 1987 or 1992 follow-up questionnaire among those who had not reported having 

asthma upon enrollment in 1977. During the 10-year follow-up (1977 to 1987), the incidence of 

new asthma was 2.1% for males and 2.2% for females (Greer et al., 1993). Ozone concentration 

data were not provided. A relative risk of 3.12 (95% CI: 1.16, 5.85) per 10 ppb increase in 

annual mean O3 (exposure metric not stated) was observed in males, compared to a relative risk 

of 0.94 (95% CI: 0.65, 1.34) in females. In the 15-year follow-up study (1977-1992), 3.2% of 

the eligible males and a slightly greater 4.3% of the eligible females developed adult asthma 

(McDonnell et al., 1999). The mean 20-year average for 8-h avg O3 (9 a.m. to 5 p.m.) was 

46.5 ppb (SD 15.3) from 1973 to 1992. For males, the relative risk of developing asthma was 

2.27 (95% CI: 1.03, 4.87) per 30 ppb increase in 8-h avg O3. Once again, there was no evidence 

of an association between O3 and new-onset asthma in females (relative risk of 0.85 [95% CI: 

0.55, 1.29]). The lack of an association does not necessarily indicate no effect of O3 on the 

development of asthma among females. For example, differences between females and males in 

time-activity patterns may influence relative exposures to O3, leading to greater misclassification 

of exposure in females. None of the other pollutants (PM10, SO4

2-, SO2, and NO2) were 

associated with development of asthma in either males or females. Adjusting for copollutants 

did not diminish the association between O3 and asthma incidence for males. The consistency of 

the results in the two studies with different follow-up times, as well as the independent and 

robust association between annual mean O3 concentrations and asthma incidence, provide 

supportive evidence that long-term O3 exposure may be associated with the development of 

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asthma in adult males. However, because the AHSMOG cohort was drawn from a narrow 

subject definition, the representativeness of this cohort to the general U.S. population may 

be limited. 

A similar study of incident asthma cases in relation to air pollutants (O3, PM, NO2, and 

acid vapors) among children was carried out in the Children’s Health Study (McConnell et al., 

2002). A total of 3,535 initially nonasthmatic children (ages 9 to 16 years at enrollment) were 

followed for up to 5 years to identify new-onset asthma cases. Communities were stratified by 

pollution levels, with six high-O3 communities (mean 1-h max O3 of 75.4 ppb [SD 6.8] over four 

years) and six low-O3 communities (mean 50.1 ppb [SD 11.0]). Ozone concentrations were not 

strongly correlated with the other pollutants. New diagnoses of asthma were reported for 

265 children during the follow-up period. Asthma risk was not higher for residents of the six 

high-O3 communities versus residents of the six low-O3 communities. However, within the 

high-O3 communities, asthma risk was 3.3 (95% CI: 1.9, 5.8) times greater for children who 

played three or more sports as compared with children who played no sports. None of the 

children who lived in high-O3 communities and played three or more sports had a family history 

of asthma. This association was absent in the low-O3 communities (relative risk of 0.8 [95% CI: 

0.4, 1.6]). No associations with asthma were seen for PM10, PM2.5, NO2, or inorganic acid 

vapors. These results suggest effect modification by physical activity of the impacts of O3 on 

asthma risk. The overall observed pattern of effects of sports participation on asthma risk was 

robust to adjustment for socioeconomic status, history of allergy, family history of asthma, 

insurance, maternal smoking, and body mass index. Playing sports may index greater outdoor 

activity when O3 levels are higher and an increased ventilation rate, which may lead to increased 

O3 exposure. It should be noted, however, that these findings were based on a small number of 

new asthma cases (n = 29 among children who played three or more sports) and were not based 

on a priori hypotheses. Future replication of these findings in other cohorts would lend greater 

weight to a causal interpretation. 

Recent cross-sectional surveys have detected no associations between long-term O3 

exposures and asthma prevalence, asthma-related symptoms, or allergy to common aeroallergens 

in children after controlling for covariates (Charpin et al., 1999; Kuo et al., 2002; Ramadour 

et al., 2000). It should be noted that O3 levels were quite low in all cases, with a range of 16 to 

27 ppb for 8-h max O3. The longitudinal study design, which observes new onset of asthma 

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prospectively, provides stronger evidence in relation to the question of O3 effects on asthma 

development. 

7.5.7 Respiratory Effects of Chronic Ozone Exposure on Susceptible 

Populations 

Studies on the effect of long-term O3 exposure on respiratory health have focused mostly 

on children, a potentially susceptible population. Seasonal O3 exposure was associated with 

smaller increases in lung function and respiratory inflammation in children. Other studies have 

investigated additional groups of potentially susceptible individuals. In the Children’s Health 

Study, McConnell et al. (1999) examined the association between O3 levels and the prevalence 

of chronic lower respiratory tract symptoms in 3,676 southern California children with asthma. 

In this cross-sectional study, bronchitis, phlegm, and cough were not associated with annual 

mean O3 concentrations in children with asthma or wheeze. All other pollutants examined 

(PM10, PM2.5, NO2, and gaseous acid) were associated with an increase in phlegm but not cough. 

The mean annual average 1-h max O3 concentration was 65.6 ppb (range 35.5 to 97.5) across the 

12 communities. 

In another Children’s Health Study analysis, McConnell et al. (2003) evaluated 

relationships between air pollutants and bronchitic symptoms among 475 children with asthma. 

The mean 4-year average 8-h avg O3 (10 a.m.-6 p.m.) concentration was 47.2 ppb (range 28.3 to 

65.8) across the 12 communities. For a 1 ppb increase in 8-h avg O3 averaged over 4 years, the 

between-community odds ratio was 0.99 (95% CI: 0.98, 1.01) compared to the withincommunity odds ratio of 1.06 (95% CI: 1.00, 1.12). The authors commented that if the larger 

within-community effect estimates were correct, then other cross-sectional (betweencommunity) studies might have underestimated the true effect of air pollution on bronchitic 

symptoms in children. These differences might be attributable to confounding by poorly 

measured or unmeasured risk factors that vary between communities. PM2.5, NO2, and organic 

carbon also were associated with bronchitic symptoms. In two-pollutant models, the withincommunity effect estimates for O3 were markedly reduced and no longer significant in some 

cases. 

One recent study examined a susceptible group not examined before. Goss et al. (2004) 

investigated the effect of O3 on pulmonary exacerbations and lung function in individuals over 

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the age of 6 years with cystic fibrosis (n = 11,484). The study included patients enrolled in the 

Cystic Fibrosis Foundation National Patient Registry, which contains demographic and clinical 

data collected annually at accredited centers for cystic fibrosis. For 1999 through 2000, the 

annual mean O3 concentration, calculated from 1-h averages from 616 monitors in the U.S. EPA 

Aerometric Information Retrieval System (AIRS), was 51.0 ppb (SD 7.3). Exposure was 

assessed by linking air pollution values from AIRS with the patient’s home ZIP code. No clear 

association was found between annual mean O3 and lung function parameters. However, a 

10 ppb increase in annual mean O3 was associated with a 10% (95% CI: 3, 17) increase in the 

odds of two or more pulmonary exacerbations. Significant excess odds of pulmonary 

exacerbations also were observed with increased annual mean PM10 and PM2.5 concentrations. 

The O3 effect was robust to adjustment for PM10 and PM2.5, 8% (95% CI: 1, 15) increase in 

odds of two or more pulmonary exacerbations per 10 ppb increase in annual mean O3. 

In summary, few studies have identified and investigated potentially susceptible 

populations. Results from the Children’s Health Study do not provide strong evidence that 

asthmatic children are particularly susceptible to long-term exposure to O3. Ozone exposure 

was, however, shown to be associated with adverse respiratory health responses in individuals 

with cystic fibrosis. 

7.5.8 Effects of Chronic Ozone Exposure on Mortality and Cancer Incidence 

PM

There is inconsistent and inconclusive evidence for a relationship between long-term O3 

exposure and increased mortality and cancer risk (see Table AX7-7 in Annex 7, Section AX7.1). 

In the Harvard Six Cities Study (Dockery et al., 1993), adjusted mortality rate ratios were 

examined in relation to long-term mean O3 concentrations in six cities: Topeka, KS; St. Louis, 

MO; Portage, WI; Harriman, TN; Steubenville, OH; and Watertown, MA. Mortality rate ratios 

were adjusted for age, sex, smoking, education, and body mass index. Mean O3 concentrations 

from 1977 to 1985 ranged from 19.7 ppb in Watertown to 28.0 ppb in Portage. Long-term 

mean O3 concentrations were not found to be associated with mortality in the six cities. 

However, strong associations were observed between long-term mean concentrations of PM15, 

2.5, and sulfate particles and mortality. 

In a large prospective cohort study of approximately 500,000 U.S. adults, Pope et al. 

(2002) examined the effects of long-term exposure to air pollutants on mortality (American 

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Cancer Society, Cancer Prevention Study II). All cause, cardiopulmonary, lung cancer, and all 

other cause mortality risk estimates for long-term O3 exposure are shown in Figure 7-27. No 

consistent positive associations were observed between O3 and mortality. The mortality risk 

estimates were larger when analyses were restricted to the summer months (July to September) 

when O3 levels were generally higher. The mean summertime 1-h max O3 level from 1982 to 

1998 was 59.7 ppb (SD 12.8). The O3-mortality risk estimates were positive for all-cause and 

cardiopulmonary mortality, with a marginally significant estimate for cardiopulmonary mortality 

in the summer months; but a nonsignificant negative O3 risk estimate was observed for lung 

cancer mortality. In contrast, consistent positive and significant effects of PM2.5 were observed 

for both lung cancer and cardiopulmonary mortality. 

Figure 7-27. Adjusted O3-mortality relative risk estimates (95% CI) by cause of mortality 

and time period of analysis per subject-weighted mean O3 concentration in 

the Cancer Prevention Study II by the American Cancer Society. 

Source: Derived from Pope et al. (2002). 

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Lipfert et al. (2000b, 2003) reported positive effects on all-cause mortality for peak O3 

exposures (95th percentile levels) in the U.S. Veterans Cohort study of approximately 50,000 

middle-aged men recruited with a diagnosis of hypertension. The actual analysis involved 

smaller subcohorts based on exposure and mortality follow-up periods. Four separate exposure 

periods were associated with three mortality follow-up periods. The mean 95th percentile O3 

concentration ranged from 85 to 140 ppb in the four exposure periods. In a preliminary 

screening of regression results, Lipfert et al. (2000b) found a negative association for mean O3 

and a positive relationship for peak O3; thus, peak O3 was used in subsequent analyses. 

The mean of the peak values ranged from 85 to 140 ppb over the four exposure periods. 

For concurrent exposure periods, peak O3 was positively associated with all-cause mortality, 

with a 9.4% (95% CI: 0.4, 18.4) excess risk per mean 95th percentile O3 less estimated 

background level (not stated). When exposure periods preceding death were considered, no 

association between O3 and mortality was observed (!0.2% [95% CI: !12.5, 12.1]). In a further 

analysis, Lipfert et al. (2003) reported the strongest positive association for concurrent exposure 

to peak O3 for the subset of subjects with low diastolic blood pressure during the 1982 to 1988 

period. Once again, the O3 effect was diminished when exposure (1982 to 1988) preceded 

mortality (1989 to 1996). 

A long-term prospective cohort study (AHSMOG) of 6,338 nonsmoking, non-Hispanic 

white individuals living in California examined the association between air pollutants and lung 

cancer incidence (Beeson et al., 1998). The mean monthly average 24-h avg O3 level from 1973 

to 1992 was 26.2 ppb (SD 7.7). Over the follow-up period of 1977 to 1992, 20 females (35% 

smokers, n = 7) and 16 males (37.5% smokers, n = 6) developed lung cancer. An association 

was observed between long-term O3 exposure and increased incidence of lung cancer in males 

only. The relative risk for incident lung cancer among males was 3.56 (95% CI: 1.35, 9.42) for 

an interquartile range increase in hours per year (556 hours/year) when O3 levels exceeded 

100 ppb. A stronger association was seen for males who never smoked (4.48 [95% CI: 1.25, 

16.04]) compared to those who smoked in the past (2.15 [95% CI: 0.42, 10.89]) (Beeson et al., 

1998). 

A related study by Abbey et al. (1999) examined the effects of long-term O3 exposure on 

all-cause (n = 1,575), cardiopulmonary (n = 1,029), nonmalignant respiratory (n = 410), and lung 

cancer (n = 30) mortality in the same AHSMOG study population. A particular strength of this 

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study was the extensive effort devoted to assessing long-term air pollution exposures, including 

interpolation to residential and work locations from monitoring sites over time and space. 

No associations with long-term O3 exposure were observed for all cause, cardiopulmonary, and 

nonmalignant respiratory mortality. However, effects of O3 on lung cancer mortality paralleled 

the results of the previous study by Beeson and colleagues. That is, an association between lung 

cancer mortality and chronic O3 exposure was observed in males only, with a relative risk of 

4.19 (95% CI: 1.81, 9.69) for an interquartile range increase in hours per year (551 hours/year) 

when O3 levels exceeded 100 ppb. The gender-specific O3 effects may be partially attributable 

to differences in activity and time spent outdoors. The questionnaires indicated that males spent 

approximately twice as much time outdoors and performed more vigorous exercises outdoors, 

especially during the summer, compared to the females. However, the very small numbers of 

lung cancer deaths (n = 12 for males) raise concerns with regard to the precision of the effect 

estimate, as evidenced by the wide confidence intervals. The lack of an association of chronic 

O3 exposure with other mortality outcomes, which had much larger samples sizes, also is of 

concern. A study by Pereira et al. (2005) provides limited evidence possibly suggestive of an 

association between O3 and increase risk of cancer. Correlations between average air pollution 

data from 1981 to 1990 and cases of larynx and lung cancer in 1997 were assessed in 

communities of São Paulo, Brazil. Of all the pollutants examined (PM10, NO2, NOx, SO2, CO, 

and O3), O3 was best correlated with cases of larynx (r = 0.99, p = 0.01) and lung (r = 0.72, 

p = 0.28) cancer. 

Only a few studies have examined the effect of chronic O3 exposure on mortality outcomes 

and incidence of cancer. Consistent associations with long-term O3 exposure have not been 

observed for all-cause and cardiopulmonary mortality. There is some very limited evidence 

suggestive of an association between chronic O3 exposure and lung cancer incidence and 

mortality; however, uncertainty remains due to the small number of available studies and the 

very small numbers of lung cancer cases observed in most of the studies. Of particular note is 

the fact that the very large American Cancer Society study did not find any association between 

long-term O3 concentrations and lung cancer. In addition, the weight of evidence from 

toxicologic studies does not support ambient O3 as a pulmonary carcinogen in laboratory animal 

models (National Toxicology Program, 1994). 

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7.5.9 Effects of Ozone on Birth-Related Health Outcomes 

In recent years, air pollution epidemiologic studies have examined impacts on birth-related 

endpoints, including intrauterine, perinatal, postneonatal, and infant deaths; premature births; 

intrauterine growth retardation; very low birth weight (weight <1500 grams) and low birth 

weight (weight <2500 grams); and birth defects. However, the majority of these studies did not 

examine the effect of O3. In those limited studies that investigated O3, no associations were 

found between O3 and birth outcomes, with the possible exception of birth defects. The 

following is a synopsis of the literature on this topic. 

Pereira et al. (1998) investigated the impacts of NO2, SO2, CO, O3, and PM10 on 

intrauterine mortality in São Paulo, Brazil during 1991 and 1992. Mean 1-h max O3 was 

34.8 ppb (SD 23.2). Intrauterine mortality was most significantly associated with NO2, and less 

for SO2 and CO. No association was found for O3 or PM10. Pereira et al. (1998) also sampled 

umbilical cord blood from healthy nonsmoking pregnant women soon after delivery in 1995 

and analyzed the blood for carboxyhemoglobin. They found an association between 

carboxyhemoglobin levels and ambient CO after adjusting for passive smoking and weight, 

suggesting an impact of CO on the fetus. Also, Loomis et al. (1999) examined the association 

between several air pollutants (NO2, SO2, CO, O3, PM2.5) and infant mortality in Mexico City in 

the years 1993 to 1995. Mean 24-h avg O3 was 44.1 ppb (SD 15.7). They reported that the 

strongest association was found for PM2.5 with a 3- to 5-day cumulative lag. They noted that 

infant mortality also was associated with NO2 and O3 at a 3- to 5-day lag, but not as consistently 

as PM2.5. There have been other studies that examined possible air pollution effects on 

postneonatal mortality (Bobak and Leon, 1992; Bobak and Leon, 1999; Kaiser et al., 2004; 

Woodruff et al., 1997), but these studies did not examine O3. 

Ritz and Yu (1999) investigated the effects of ambient CO on low birth weight among 

children born in southern California between 1989 and 1993. They focused on CO, stating that 

“a biologic mechanism for fetal effects has been proposed for CO, but not for other air 

pollutants.” They found that exposure to higher levels of ambient CO during the last trimester 

was associated with an increased risk for low birth weight. Using available data, they also 

estimated last-trimester exposures for NO2, O3, and PM10. The last trimester average O3 level 

was 20.9 ppb (range 3.0 to 49.4). NO2 and PM10 were positively associated with CO, but O3 was 

negatively associated with CO (r = !0.65). Ha et al. (2001) examined CO, NO2, SO2, O3, and 

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TSP for their associations with low birth weight in Seoul, Korea during the years 1996 and 1997 

(data were analyzed using Poisson GAM with default convergence parameters). They estimated 

first and third trimester exposures by averaging daily air pollution levels during the 

corresponding days for the registered births. The median 8-h avg O3 levels were 22.4 ppb 

(IQR 13.6) for the first trimester and 23.3 ppb (IQR 16.1) for the third trimester. Ha et al. (2001) 

found that first trimester exposures of CO, NO2, SO2, and TSP were associated with increased 

risk of low birth weight, whereas O3 was associated with a decreased risk. The opposite pattern 

was observed for third trimester exposures, with an increased risk of low birth weight found only 

for O3. When exposures from both trimesters were examined simultaneously, the associations of 

first trimester exposures of CO, NO2, SO2, and TSP with increased risk of low birth weight 

remained; however, the association between third trimester O3 exposure and low birth weight 

was diminished. Based on these results, Ha et al. concluded that exposures to CO, NO2, SO2, 

and TSP in the first trimester were risk factors for low birth weight. Note that neither of these 

studies examined the air pollution effect by season. Other studies that examined the associations 

between air pollution and low birth weight (Bobak, 2000; Bobak and Leon, 1999; Lin et al., 

2001; Maisonet et al., 2001; Wang et al., 1997) found associations between low birth weight and 

either one or more of CO, SO2, NO2 and PM indices, but did not examine O3 data. Collectively, 

these results do not provide any credible evidence of O3 contributing to low birth weight. 

Two studies by Dejmek et al. (1999, 2000) examined the relationship between ambient air 

pollution and risk of intrauterine growth retardation in a highly polluted area of Northern 

Bohemia (Teplice District). Both studies, however, focused on PM indices and did not analyze 

gaseous pollutants, such as O3. 

A few studies have examined the association between air pollution and premature births 

(Bobak, 2000; Ritz et al., 2000; Xu et al., 1995), but only Ritz et al. (2000) included O3 in their 

analysis. Ritz et al. evaluated the effect of air pollution exposure during pregnancy on the 

occurrence of preterm birth in a cohort of 97,518 neonates born in southern California. 

CO, NO2, SO2, O3, and PM10 data measured at 17 air quality monitoring stations were used to 

estimate the average exposures for the first month and the last 6 weeks of pregnancy. The 

mean 8-h avg O3 (9 a.m. to 5 p.m.) level was 36.9 ppb during both time periods. They found 

associations between PM10 levels averaged for the last 6 weeks of pregnancy as well as PM10 

levels averaged over the first month of pregnancy. Similar but weaker associations were found 

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for CO, but no association was found for O3. The reported correlation matrix indicated that O3 

was negatively correlated with CO (r = !0.45) and only weakly correlated with PM10 (r = 0.2). 

Other results from Beijing, China (Xu et al., 1995) and the Czech Republic (Bobak, 2000) 

suggested that SO2 and TSP were associated with preterm births. Considering that O3 tends to be 

negatively correlated with winter-type pollutants, O3 is unlikely to be an important risk factor for 

preterm births. 

Ritz et al. (2002) evaluated the effect of air pollution on the occurrence of birth defects in 

neonates and fetuses delivered in southern California from 1987 to 1993 as ascertained by the 

California Birth Defects Monitoring Program. They averaged air pollution (CO, O3, PM10, 

and NO2) levels measured at the assigned ambient station over the first, second, and third month 

of gestation. Conventional, polytomous, and hierarchical logistic regressions were used to 

estimate odds ratios for subgroups of cardiac and orofacial defects. Concentration-response 

relationships were observed for second month CO exposure on ventricular septal defects, and 

second month O3 exposure on aortic artery and valve defects, pulmonary artery and valve 

anomalies, and conotruncal defects. The odds ratios observed for these outcomes were similar 

and quite large (e.g., the odds ratios comparing the highest [monthly 24-h avg mean 34.9 ppb] 

to lowest [monthly mean 6.4 ppb] O3 quartiles ranged from 2.0 to 2.7), and were not sensitive in 

multipollutant models. Ritz et al. (2002) reported that they did not observe consistently 

increased risks and concentration-response patterns for NO2 and PM10 after controlling for the 

effects of CO and O3. Results from this study contrast to those for other birth-related outcomes, 

in that both CO and O3 (presumably negatively correlated pollutants) were associated with birth 

defects. Further, O3 showed associations with more birth defect outcomes than did CO. 

It should be noted, however, that the concentration-response relationships were quite specific to 

exposures during the second month. Associations with third month exposures were often 

negative (though not significantly). Since both CO and O3 show strong seasonal peaks, it is 

possible that seasonal confounding could have played some role in these associations. This is 

the only study to date that examined the relationship between air pollution and birth defects. 

In summary, O3 was not an important predictor of several birth-related outcomes including 

intrauterine and infant mortality, premature births, and low birth weight. Birth-related outcomes 

generally appear to be associated with air pollutants that tend to peak in the winter and are 

possibly traffic-related, most notably CO. The strong results for CO are consistent with its 

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ability to cross the placental barrier and the high affinity that hemoglobin in fetal blood has for 

binding with it. However, given that most of these studies did not analyze the data by season, 

seasonal confounding may have therefore influenced the reported associations. One study 

reported some results suggestive of associations between exposures to O3 in the second month of 

pregnancy and birth defects, but further evaluation of such potential associations is needed. 

7.5.10 Summary of Chronic Ozone Exposure Effects on Morbidity 

and Mortality 

C In the past decade, important new longitudinal studies have examined the effects of 

chronic O3 exposure on respiratory health outcomes, including seasonal declines in 

lung function, increases in inflammation, and development of asthma in children and 

adults. Seasonal O3 effects on lung function have been reported in several studies. 

In contrast to supportive evidence from chronic animal studies, epidemiologic studies 

of inflammation, new asthma development, and longer-term lung function declines 

remain inconclusive. 

C Few studies have investigated the effect of long-term O3 exposure on mortality and 

cancer incidence. Uncertainties regarding the exposure period of relevance, 

differential effects by gender, and inconsistencies across outcomes raise concerns 

regarding plausibility. There is currently little evidence for a relationship between 

chronic O3 exposure and increased risk of lung cancer or of mortality. 

C A limited number of studies have examined the relationship between air pollution and 

birth-related health outcomes, including mortality, premature births, low birth weights, 

and birth defects. The most consistent associations with various birth outcomes were 

observed for CO, with very little credible evidence being found for any O3 effects. 

7.6 INTERPRETIVE ASSESSMENT OF THE EVIDENCE IN 

EPIDEMIOLOGIC STUDIES OF OZONE HEALTH EFFECTS 

7.6.1 Introduction 

In the 1996 O3 AQCD, discussion of the available epidemiologic information focused 

primarily on individual-level camp and exercise studies and on studies of hospital admissions 

and emergency room visits. The field studies indicated concentration-response relationships of 

ambient air O3 exposure with declines in pulmonary function, increases in respiratory symptoms, 

and exacerbation of asthma, especially in children. Numerous new studies provide additional 

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evidence for evaluating associations between O3 exposure and the above respiratory health 

outcomes. The 1996 O3 AQCD review of aggregate population time-series studies indicated an 

association between ambient O3 concentrations and increased hospitalizations. A limited 

number of studies examined O3-mortality relationships. The current O3 AQCD presents results 

from time-series studies that have addressed previously unresolved issues regarding potential 

linkages between ambient O3 concentrations and health outcomes, particularly mortality. Daily 

time-series studies minimize confounding by population characteristics (e.g., cigarette smoking, 

diet, occupation) by following the same population from day to day. However, confounders 

operating over shorter time scales can affect O3 risk estimates in these studies. 

In this section, issues and attendant uncertainties that affect the interpretation of available 

epidemiologic evidence regarding O3 health effects are discussed. The use of various indices to 

represent O3 exposure in epidemiologic studies is discussed first. Also, of interest is the issue of 

confounding by temporal factors, meteorological factors, and copollutants. The shape of the 

concentration-response function and heterogeneity of O3 effects are also discussed briefly. 

All of these topics are important for characterizing and interpreting ambient O3-health effects 

associations. 

7.6.2 Ozone Exposure Indices 

Three O3 indices were most often used to indicate daily O3 exposure: maximum 1-h 

average (1-h max); maximum 8-h average (8-h max); and 24-h average (24-h avg). The 

8-h max O3 concentration is a frequently used index in newer epidemiologic studies, as it relates 

most closely to the averaging time of the current U.S. EPA NAAQS. The O3 exposure indices 

are highly correlated, as indicated in several studies. For example, in the 21 European 

multicities acute mortality study (Gryparis et al., 2004), 1-h max O3 was found to be highly 

correlated with 8-h max O3, with a median correlation coefficient of 0.98 (range 0.91 to 0.99). 

Among single-city studies, the 1-h max O3 and 8-h max O3 also were found to have correlation 

coefficients ranging from 0.91 to 0.99 in various areas, such as Atlanta, GA (Tolbert et al., 2000; 

White et al., 1994); southern New England (Gent et al., 2003); Ontario, Canada (Burnett et al., 

1994); and Mexico City (Loomis et al., 1996; Romieu et al., 1995). In addition, 1-h max O3 was 

highly correlated with 24-h avg O3, as observed in the Mexico City study by Loomis et al. (1996) 

(r = 0.77) and in the Ontario, Canada study by Burnett et al. (1994) (r = 0.87). 

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All studies discussed in Sections 7.2 to 7.5 were examined for presentation of the three O3 

exposure indices. Several presented the concentration data and correlations among 1-h max, 8-h 

max, and 24-h avg O3 ambient measures. Some presented the associated risk estimates of 

comparable analyses for the three exposure indices. No papers provided a statistical analysis 

comparing results from the different indices. The increments used in this document to 

standardize expressions of excess risks are 40 ppb for 1-h max O3, 30 ppb for 8-h max O3, and 

20 ppb for 24-h avg O3, as discussed in Section 7.1.3.2. Key findings derived from the assessed 

studies are discussed below, starting with two multicity mortality studies. 

In the large U.S. 95 communities study by Bell et al. (2004), increases in O3-associated 

daily mortality were estimated using all three O3 indices. The mean 24 h avg O3 level was 

appropriately 26 ppb across the 95 communities. Ozone concentrations for 1-h max O3 and 8-h 

max O3 were not provided. For the above-noted increments, the effect estimates calculated by 

Bell et al. (2004) using all available data were 1.34% (95% PI: 0.84, 1.85), 1.28% (95% PI: 

0.88, 1.73), and 1.04% (95% PI: 0.54, 1.55) excess risk in mortality for 1-h max O3, 8-h max O3, 

and 24-avg O3, respectively. A statistical test examining differences among these risk estimates 

indicated that there were no significant differences by exposure index. In the European study of 

21 cities (two of the 23 cities evaluated did not have 8-h max O3 data), the O3-mortality effect 

estimate for the summer season was slightly smaller for the 8-h max O3, 1.82% (95% CI: 0.99, 

3.06) excess risk compared to the 1-h max O3, 2.59% (95% CI: 1.32, 4.10) excess risk; however, 

the two risk estimates were not significantly different (Gryparis et al., 2004). The median 1-h 

max O3 levels ranged from 44 to 117 ppb and the median 8-h max O3 levels ranged from 30 to 

99 ppb across all cities during the summer. 

Several single-city mortality studies examined multiple O3 exposure indices (Anderson 

et al., 1996; Dab et al., 1996; Sunyer et al., 2002; Zmirou et al., 1996; Borja-Aburto et al., 1997). 

These studies did not differentiate risk estimates by exposure index, because the results were 

considered to be similar. Hospital admission studies also provided limited data for O3 index 

comparisons. Schouten et al. (1996) found similar O3 effects on total respiratory hospitalizations 

from 8-h max O3 and 1-h max O3 in the summer. Both indices resulted in a 4.0% excess risk per 

standardized increment. For emergency department visits, the examples of Delfino et al. 

(1998b) and Weisel et al. (2002) indicated no difference in effect estimate when using various O3 

indices. Tolbert et al. (2000) noted an increase in emergency room visits of 4.0% per standard 

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deviation increase (approximately 20 ppb) for both 1-h max O3 and 8-h max O3 as being 

expected, since the correlation between the indices was 0.99. 

Peak flow asthma panel studies generally used only one index; thus, there were limited 

data available for comparison. One respiratory symptom study (Gent et al., 2003) examined 

both 1-h max O3 and 8-h max O3 but noted no differences in the results. Only one FEV1 panel 

study examined more than one O3 exposure index. Chen et al. (1999) examined 1-h max O3 and 

24-h avg O3 and reported a decrement in FEV1 of !25.6 mL (95% CI: !49.1, !2.1) per 40 ppb 

increase in 1-h max O3 and !13.6 mL (95% CI: !33.2, 6.0) per 20 ppb increase in 24-h avg O3 

in children at a 1-day lag. For 2- and 7-day lags, smaller differences were observed between the 

two indices. The effect estimates calculated using 1-h max O3 and 24-h avg O3 concentrations 

were not found to be significantly different for any of the lags examined. 

Limited information is available by which to draw conclusions for comparison across the 

three O3 indices of 1-h max O3, 8-h max O3, and 24-h avg O3. Studies conducted in various 

cities have observed very high correlations among the daily O3 indices. For the same 

distributional increment, the excess health risk estimates and significance of associations were 

generally comparable for the three O3 indices across all outcomes. The high correlation among 

the indices presents a challenge in distinguishing the most appropriate measure for 

epidemiologic studies. Exploratory analyses using various O3 exposure indices are valuable in 

understanding relationships. However, to address the issue of multiple hypothesis testing, 

hypotheses that are confirmatory and exploratory should be decided a priori and reported 

accordingly. 

7.6.3 Confounding by Temporal Trends and Meteorologic Effects in 

Time-Series Studies 

The challenge of analyzing acute O3 effects in time-series studies is to avoid bias due to 

confounding by daily to seasonal temporal factors. On a seasonal scale, the analysis must 

remove the influence of the strong seasonal cycles that usually exist in both health outcomes 

and O3. On a daily scale, weather factors and other air pollutants also may confound the 

association of interest. This section discusses the interpretation of effect estimates after 

adjusting for temporal trends and meteorologic effects. 

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7.6.3.1 Assessment of Ozone Effects after Adjusting for Temporal Trends and 

Meteorologic Effects 

The relationship between O3 and health outcomes are significantly affected by temporal 

trends and meteorological factors, namely temperature. Analyses of the association between 

health outcomes and O3 concentrations using raw data, therefore, can be misleading. In Díaz 

et al. (1999), a U-shaped relationship was observed between mortality and O3 concentrations, 

and the negative portion of the slope was likely due to the opposing seasonal cycles in mortality 

(high in winter) and temperature (low in winter). 

Keatinge and Donaldson (2005) used two methods to identify confounding due to 

meteorological factors in their study of the mortality effect of air pollutants on adults aged 

65 years or older during the warm season in Greater London. One involved GAM analyses 

controlling for up to seven weather covariates, and the other considered graphical comparisons 

to show interrelations of mortality with weather factors and pollutants. Both methods 

pre-adjusted for weather variables, which is a more stringent test of the air pollution effect as 

compared to co-adjustment. Few of the individual analyses found a significant O3 effect, which 

tended to be small in comparison to the weather effect. Therefore, Keatinge and Donaldson 

(2005) concluded that, although air pollution may have a short-term adverse effect on health in 

London during the warm season, the results suggested that measures to prevent excess mortality 

in hot weather should focus on control of heat stress. Goldberg and Burnett (2003) reported a 

positive slope for the temperature-mortality relationship being fitted most tightly in the mild 

temperature range where mortality effects of temperature were not expected. While it is possible 

that temperature has mortality effects in the mild temperature range, daily fluctuations of air 

pollutants (especially O3) are strongly influenced by weather conditions and ascribing the 

association between temperature and mortality entirely to effects of temperature may 

underestimate the effects of air pollution. 

Sensitivity analyses to examine confounding of temperature on the O3 effect were 

performed in the U.S. 95 communities study by Bell et al. (2004). In analyses excluding days 

with high temperature (that used a cutoff as low as 29 °C [85 °F]), the range of estimated effect 

sizes (1.00 to 1.10% excess risk in mortality per 20 ppb increase in 24-h avg O3) did not differ 

from the effect estimate using all the data (1.04% excess risk). Effect estimates were found to be 

slightly higher at lower temperatures. In a related study of 19 U.S. cities, Huang et al. (2005) 

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examined the potential influence of heat waves, modeled as a natural cubic spline of the 

interaction between current temperature and average temperature in the three previous days, on 

the O3 effect on cardiopulmonary mortality. The effect estimate was generally robust to 

inclusion of this interaction term, with only a slight decrease in the estimate from 2.52% (95% 

PI: 0.94, 4.10) to 2.23% (95% PI: 0.76, 3.75) excess risk per 20 ppb increase in 24-h avg O3. 

Bell et al. (2004) and Huang et al. (2005) also found that O3 effects were robust to the 

selection of degrees of freedom for smoothing of temporal trends. Bell et al. observed that 

changing the degrees of freedom from 7 to 21 per year did not significantly affect the 

O3-mortality estimates, with effect estimates ranging from 0.82 to 1.08% excess risk per 20 ppb 

increase in 24-h avg O3 during the previous week. Huang et al. (2005) examined the sensitivity 

of summertime O3 risk estimates to varying degrees of freedom from 4 to 16 per year. The 

extent of change in the risk estimates, though varied from city to city (graphically presented), 

was not substantial. Using more degrees of freedom in temporal trend fitting (i.e., controlling 

shorter temporal fluctuations) means ascribing more details of daily health outcomes to 

unmeasured potential confounders and possibly weakening real weather and air pollution effects. 

However, results from these large multicity studies indicated that O3 effects were robust to 

aggressive smoothing of temporal trends. 

Ito et al. (2005) examined sensitivity of O3-mortality risk estimates to the extent of 

temporal trend adjustment and to alternative weather model specifications using data from seven 

U.S. cities (Cook County, IL; Detroit, MI; Houston, TX; Minneapolis, MN; New York City; 

Philadelphia, PA; and St. Louis, MO). They found that varying the degrees of freedom from 4 to 

26 per year did not substantially or systematically affect the O3-mortality estimates, except for 

Cook County where the percent excess O3-mortality risk estimates were considerably reduced 

when the temporal adjustment term with 26 degrees of freedom was applied. Ito et al. (2005) 

noted that the O3 risk estimates were generally more sensitive to alternative weather models 

(up to a 2-fold difference in percent excess risk) than to the degrees of freedom for temporal 

adjustment. Of the four weather models examined, the four-smoother model similar to the one 

used in the NMMAPS studies (Bell et al., 2004; Huang et al., 2005) generally resulted in the 

smallest effect estimates. 

Schwartz (2005) examined the sensitivity of the O3-mortality relationship to methods used 

to control for temperature. Initially, temperature lagged 0 and 1 day was controlled using 

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nonlinear regression splines with 3 degrees of freedom each. In a comparison analysis, control 

days were restricted to a temperature-matched subset. The effect estimates for all-year data 

using nonlinear regression splines (0.8% [95% CI: 0.1, 1.4] excess risk per 40 ppb increase in 

1-h max O3) and temperature-matched controls (0.9% [95% CI: 0.04, 1.8] excess risk) were not 

significantly different. Results were similar when restricting analysis to warm-season-only data. 

Temporal cycles in daily hospital admissions or emergency department visits are often 

considerably more episodic and variable than is usually the case for daily mortality. As a result, 

smoothing functions that have been developed and tuned for analyses of daily mortality data 

may not work as well at removing cyclic patterns from morbidity counts. Two methods are 

commonly used to adjust for temporal trends. The pre-adjustment method involves applying the 

adjustment to both outcome and air pollution variables prior to the regression analysis. The 

co-adjustment method involves applying the adjustment as part of the regression analysis, by 

fitting a function of time while simultaneously fitting the regression effect of air pollution and 

weather factors. As shown in a hospital admissions study by Burnett et al. (2001), the 

co-adjustment approach may lead to biased air pollution effect estimates in cases where both 

outcome and pollution variables exhibit strong seasonal cycles (note that the data were analyzed 

using Poisson GAM with default convergence criteria). Using year-round data, pre-adjustment 

followed by regression analysis yielded a 14% (95% CI: 5, 24) increase in admissions per 

40 ppb increase in 1-h max O3 with a multiday lag of 0 to 4 days. The co-adjustment method 

resulted in a 7% (95% CI: 3, 11) decrease in admissions. When the authors limited the analysis 

to the warm season (May-August), both methods yielded similar results (32% [95% CI: 21, 44] 

versus 30% [95% CI: 17, 45] increase for co-adjustment and pre-adjustment, respectively), 

implying that stratification by season can remove a significant amount of the confounding 

seasonality (which also may include seasonally-varying population behavior and ventilation 

conditions). This finding may be important to consider in reviewing the acute O3 mortality and 

morbidity literature, because the vast majority of studies published over the past decade have 

used the co-adjustment method. However, the use of pre-adjustment versus co-adjustment in 

time-series studies is an unresolved issue. More empirical research in different locales is needed 

to evaluate the merits of these two methods as far as O3 is concerned and to determine what 

endpoints may be affected. 

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More sensitivity analysis of O3 effect estimates to the extent of adjustment for temporal 

trends and meteorological factors is needed, but perhaps it is equally as important to evaluate the 

epidemiologic adequacy of a given adjustment. For example, do the fitted mortality series 

sufficiently depict influenza epidemics? Or, when larger degrees of freedom (e.g., 12 degrees of 

freedom per year) are used, what “unmeasured” confounders, other than weather and pollution, 

are the investigators trying to adjust? Even in PM studies that conducted sensitivity analyses, 

investigators rarely stated assumptions clearly or provided adequate discussions as to potential 

reasons for the sensitivity of results. 

Given their relationship to health outcomes and O3 exposure, adjusting for temporal trends 

and meteorologic factors is critical to obtain meaningful O3 effect estimates. While the 

prevailing analytical approaches fit the data flexibly, the estimated effects of meteorologic 

variables and their impact on the adjusted O3 effects are not adequately discussed. More work is 

needed in this area to reduce the uncertainty involved in the epidemiologic interpretation of O3 

effect estimates. 

7.6.3.2 Importance of Season-Specific Estimates of Ozone Health Effects 

Analysis of O3 health effects is further complicated, as relationships of O3 with other 

pollutants and with temperature appear to change across seasons. Moolgavkar et al. (1995) 

examined the relationship between daily mortality and air pollution by season in Philadelphia, 

PA for the period of 1973 to 1988. During the summer, there was a positive relationship 

between O3 and TSP, as well as O3 and SO2. In contrast, the relationship of O3 with TSP 

and SO2 was inverse during the winter. Ozone showed positive associations only in the summer 

when the mean O3 concentration was the highest. The effect of O3 on mortality was negative 

(though not significantly) in the winter when the mean O3 concentration was low. In the summer 

multipollutant model, O3 was the only pollutant that remained significant. Similar results were 

found in another Philadelphia study by Moolgavkar and Luebeck (1996). Neither studies 

analyzed year-round data; therefore, the relationship between the excess risk estimates for all 

year and each season could not be compared. The results from these studies, however, suggest 

that year-round analyses may mask associations (positive or negative) that may exist in 

particular seasons. 

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Ito et al. (2005) examined O3-mortality associations in seven U.S. cities, but also described 

the relationship between O3 and PM for summer months (June through August) and winter 

months (December through February) in these cities (see Figure 7-28). The O3-PM relationships 

were positive in the summer and negative in the winter in all of these cities, except in Houston, 

where the O3-PM association was not clearly positive in the warmer months but positive in 

colder months. Ito et al. (2005) found that O3-mortality associations were mostly weaker, null, 

or even negative in the winter compared to the summer in most of these cities. Once again, the 

exception was Houston where the cold season O3-mortality association was positive and larger 

than those for year-round or warmer months. Findings from this study suggest the influence of 

seasonal O3-PM relationships on O3-mortality associations. It should be noted that the contrasts 

in O3 risk estimates between warm and cold months varied across the alternative weather models 

examined, with the four-smoother model generally showing the least contrast. 

In the analyses of the U.S. 90 cities data (of which 80 cities had O3 data available) by 

Samet et al. (2000; reanalysis Dominici et al., 2003), the focus of the study was PM10, but O3 and 

other gaseous pollutants also were analyzed in single- and multiple-pollutant models. In the 

reanalysis (Dominici et al., 2003), O3 was associated with an excess risk of mortality in analyses 

of all available data (0.4% [95% PI: 0.1, 0.7] excess risk per 20 ppb increase in 24-h avg O3 at a 

1-day lag) and summer only data (1.0% [95% PI: 0.5, 1.6]; however, a negative association was 

observed for the winter only analysis (!1.1% [95% PI: !2.2, 0.1]). A 2-fold greater effect was 

estimated using summer data compared to all available data. The mean 24-h avg O3 levels 

using all available data ranged from 12 to 36 ppb across the 80 cities. Season-specific O3 

concentrations were not presented. It should be noted that the analyses by Samet et al. and 

Dominici et al. used a weather model specification that is more detailed than other studies in that 

it had multiple terms for temperature and dewpoint (these two variables are generally highly 

correlated). Thus, it is possible that the high concurvity of O3 with these weather covariates may 

have produced these conflicting results. Another possibility is that the apparent negative 

relationship between O3 and mortality in the winter may have been due to confounding by PM. 

In the larger U.S. 95 communities study by Bell et al. (2004), the all-available data and warmseason-only analyses also indicated positive risk estimates (1.04% [95% PI: 0.54, 1.55] and 

0.78% [95% PI: 0.26, 1.30] excess risk per 20 ppb increase in 24-h avg O3, respectively, using a 

constrained distributed 7-day lag model), but the two estimates were similar in magnitude. 

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Figure 7-28. The relationship between PM and O3 in the summer (June through August) 

and the winter (December through February) as sorted and averaged by 

quintiles of PM. 

Source: Derived from Ito et al. (2005). 

The median 24-h avg O3 levels ranged from 14.4 to 37.3 ppb in the 55 communities with all-year 

data and from 20.4 to 36.2 ppb in the 40 communities with warm-season only data. Cold-season 

only analyses were not performed. Results from the U.S. 95 communities study appear to 

conflict with the strong seasonal variation observed in the U.S. 90 cities study. However, there 

are several differences between the two studies that might account for these results. First, the 

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U.S. 95 communities study nearly doubled the study period by extending the analysis by six 

additional years (1987 to 2000 versus 1987 to 1994) and it included 15 additional cities in 

addition to the original 80. Also, the warm seasons are characterized differently in the two 

studies. The U.S. 90 cities study defined summer as a three-month period of June through 

August, whereas the 95 communities study defined the warm season as a 7-month period of 

April through October. In addition, the results presented in the U.S. 90 cities study were from a 

single-day lag model (1-day lag) while the estimates from the 95 communities study were 

calculated using a constrained distributed 7-day lag model. Differences in seasonal O3 effects 

observed in the two related studies might be attributable to some of these factors. 

Many studies reported larger excess mortality risks in the warm (or summer) season than in 

the cool (or winter) season (see Figure 7-21 in Section 7.4.5). These studies showed cool-season 

risk estimates that were either smaller compared to warm-season estimates or slightly negative. 

Of the studies that analyzed data by season, only one study in Pittsburgh, PA (Chock et al., 

2000) showed negative risk estimates for the summer. Ozone concentration data were not 

provided for the Chock et al. (2000) study. The studies that observed larger, positive 

associations between O3 and mortality in warm seasons are consistent with the expectation 

that O3, if harmful, should have a stronger association with health outcomes for the summer 

when exposure to O3 is higher. However, the negative O3-mortality associations seen for the 

winter suggest that further examination of this issue is required. Specifically, if O3 levels in the 

winter are shown to be negatively associated with factors (e.g., PM) that are positively 

associated with mortality, then these potentially spurious negative O3-mortality associations can 

be explained. Several examples of this phenomenon also exist in morbidity studies investigating 

the effect of O3 on daily hospital admissions and emergency department visits (Anderson et al., 

1998; Burnett et al., 2001; Prescott et al., 1998; Thompson et al., 2001). 

Unlike the time-series studies examining outcomes of mortality, hospital admissions, and 

emergency department visits, most acute field studies did not perform year-round analyses. 

The acute field studies that examined the relationship between O3 and lung function, respiratory 

symptoms, and inflammation focused primarily on the O3 effect during the warm season 

when O3 levels were expected to be high and subjects spent more time outdoors and were 

physically active. 

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There are seasonal (e.g., air conditioning use) or seasonally-modified (e.g., time spent 

outdoors, air exchange rates) factors that affect the relationship between ambient concentrations 

and personal exposures to O3, as discussed in Section 3.9. The influence of combinations of 

these factors across seasons on air pollution health effects can become quite complex. For 

example, longer time spent outdoors in the summer may increase personal exposure to O3 for 

some segment of the population, but the increased use of air conditioning may reduce exposures 

to ambient O3 for those who spend much of their time indoors. In the meta-analysis by Levy 

et al. (2005), the combined risk estimate from the warm season was greater (3.38% [95% CI: 

2.27, 4.42] per 40 ppb increase in 1-h max O3) compared to the estimate from all-year data 

(1.64% [95% CI: 1.25, 2.03]). However, further analysis suggested that the O3-mortality risk 

estimates were smaller in cities with high air conditioning prevalence. Seasonal factors such as 

these that influence the relationship between ambient concentrations and personal exposures 

make the interpretation of the concentration-response relationships obtained from analyses of 

year-round data less straightforward. 

In some cities, O3 is only monitored during the warm season. For example, 34% of the 

communities in the U.S. 95 communities study only collected O3 data during the warm season 

(Bell et al., 2004). The cities with larger populations and/or higher O3 concentrations generally 

collected year-round data. There is some concern that differential data availability may also 

contribute to the seasonal differences in O3 health effects observed. 

The potential influence of season on O3 effect estimates was examined here using summary 

density curves. The O3 effect observed in all-year data was compared to effects from warmseason and cool-season only data (Figures 7-29 and 7-30). Summary probability density curves 

were calculated to assess the effect estimates from the various studies (see Annex Section AX7.2 

for further explanation of summary density curves). The summary density curves shown in 

Figures 7-29 and 7-30 were smoothed by multiplying a constant to the standard error of each 

effect estimate in the calculation of the individual distribution functions. Since the normal 

distribution is unimodal, this constant will oversmooth when the density is multimodal. 

In Figure 7-29, the summary density curves of O3-associated all-cause (nonaccidental) mortality 

are presented (see Figure 7-21 in Section 7.4.5 for the effect estimates). The summary density 

curves are calculated using results from 14 studies that reported at least two of the three 

estimates. This figure indicates that 75% of the area under the density curve has a value greater 

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All year Warm season Cool season 

% area under the density curve and >0 75% 78% 62% 

Mean (SD) effect estimates 1.2% (2.8%) 1.3% (2.6%) 0.1% (3.1%) 

Mode effect estimates 0.7% 0.9% 0.4% 

Figure 7-29. Summary density curves of the percent change in all cause mortality for 

all-year data and by season per standardized increment (see Section 7.1.3.2). 

Effect estimates from 14 studies have been included in the summary density 

curves (see Figure 7-21 in Section 7.4.5 for the effect estimates). 

than zero for all-year data compared to 78% for warm-season data and 62% for cool-season data. 

Therefore, both all-year and warm-season data generally indicate a positive O3 effect on 

mortality. The mean effect estimates for all-year data and warm-season only data are 1.2% 

(SD 2.8) and 1.3% (SD 2.6) excess risk in mortality per 40 ppb increase in 1-h max O3, 

respectively. A slightly larger mode of effects is observed for warm-season data (0.9% excess 

risk) compared to all-year data (0.7%). The cool-season-only data indicate that there is no 

excess risk (mean 0.1% [SD 3.1]) associated with O3 concentrations. 

Similar observations are made when examining the O3 effect on total respiratory hospital 

admissions (Figure 7-30). Six studies provided season-specific estimates as well as all-year 

results (see Figure 7-10 in Section 7.3.3 for the effect estimates). Once again, a large percent 

of the area under the summary density curve is greater than zero when using all-year and 

warm-season data, 92% and 84%, respectively, compared to cool-season data, 49%. 

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All year Warm season Cool season 

% area under the density curve and >0 92% 84% 49% 

Mean (SD) effect estimates 6.5% (6.4%) 6.3% (9.1%) 0.8% (6.1%) 

Mode effect estimates 1.8% 4.0% !1.3% 

Figure 7-30. Summary density curves of the percent change in total respiratory hospital 

admissions for all-year data and by season per standardized increment 

(see Section 7.1.3.2). Effect estimates from six studies have been included 

in the summary density curves (see Figure 7-10 in Section 7.3.3 for the 

effect estimates). 

The mean O3 effect estimates for warm-season-only data, 6.3% (SD 9.1) excess risk per 40 ppb 

increase in 1-h max O3, and all-year analyses, 6.5% (SD 6.4) excess risk, are similar. A larger 

mode of effects is observed for warm-season data (3.9% excess risk) compared to all-year data 

(1.8% excess risk). A small O3 effect (0.8% [SD 6.1] excess risk) is observed when using coolseason data only. 

Integrating seasonal influences across the various health outcomes supports the view 

that O3 effects are different in the cool and warm seasons, with greater effects observed during 

the warm season. Warm-season data should be used to derive quantitative relationships for the 

effect of O3 on health outcomes, because the attenuation of O3 exposure from ambient 

concentrations is likely to be much less during this season. However, studying summer data 

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only when all-year data are available weakens the power of the study, since less data are 

analyzed. In addition, increased adverse health outcomes are observed in the winter, some of 

which may be attributable to O3. The O3 effect in the wintertime may be masked by the effects 

of PM due to the negative correlation between these two variables (see Section 7.6.4.2 for 

further discussion). Therefore, analysis of all-year data may be improved by adjusting for PM 

indices in addition to adequate adjustment of meteorological factors and temporal trends. 

Seasonality influences the relationship between O3 and health outcomes, as it may serve as 

an indicator for time-varying factors, such as temperature, copollutant concentrations, 

infiltration, and human activity patterns. Given the potentially significant effect of season, O3 

effect estimates computed for year-round data need to be interpreted with caution. Finding small 

or no effects may simply reflect (a) the cancellation of positive associations in the summer and 

negative associations in the winter or (b) the presence of confounding due to the strong seasonal 

character of O3 concentrations. 

7.6.4 Assessment of Confounding by Copollutants 

Potential confounding by daily variations in copollutants is another analytical issue to be 

considered. With respect to copollutants, daily variations in O3 tend to not correlate highly with 

most other criteria pollutants (e.g., CO, NO2, SO2, PM10), but they may be more correlated with 

secondary fine PM (e.g., PM2.5, sulfates) measured during the summer months. Assessing the 

independent health effects of two pollutants that are correlated over time is not straightforward. 

If high correlations between O3 and PM or other gaseous pollutants exist in a given area, then 

disentangling their relative individual contributions to observed health effects associations 

becomes very difficult. The changing relationship between O3 and other copollutants also is at 

issue. Ito et al. (2005) described the relationship between O3 and PM by season in seven urban 

U.S. cities: Cook County, IL; Detroit, MI; Houston, TX; Minneapolis-St. Paul, MN; New York 

City; Philadelphia, PA; and St. Louis, MO (see Figure 7-28). With the exception of Houston, 

TX, the O3-PM associations were positive in the summer and negative in the winter. 

Relationships between O3 and copollutants and the potential confounding of the O3 effect by 

copollutants are discussed in the following section. 

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7.6.4.1 Relationship between Personal Exposure to Ozone and Copollutants 

The correlation between ambient O3 concentrations and ambient levels of PM, NO2, SO2, 

and CO measured at central monitoring sites has been reported in many studies; however, only a 

limited number of studies have examined the association between personal O3 exposures and 

personal exposures to other copollutants. The relationship between personal exposure to O3 and 

personal exposure to PM2.5 was examined by Chang et al. (2000) in various microenvironments 

in Baltimore, MD, using activity data from the National Human Activity Pattern Survey study 

(Klepeis, 1999). Activities were scripted to simulate activities performed by older adults 

(65+ years of age). Mean hourly personal O3 exposures were 15.0 ppb (SD 18.3) in the summer 

and 3.6 ppb (SD 7.5) in the winter. Strong correlations were not observed between 1-h personal 

O3 and PM2.5 concentrations in the various microenvironments, including the indoor residence 

(n = 91), other indoor environments (n = 53), outdoor near the roadways (n = 21), outdoor away 

from roads (n = 19), and in vehicles (n = 71). Spearman r ranged from !0.14 to 0.29 during the 

summer and !0.28 to 0.05 during the winter (SEs not provided, all correlation coefficients noted 

as statistically nonsignificant [p > 0.05]). 

An issue of particular interest is the correlation between personal exposure to O3 and 

personal exposure to the ambient component of PM2.5. In a study of susceptible populations 

(older adults [n = 20], individuals with COPD [n = 15], and children [n = 21]) in Baltimore, MD, 

a total of 800 person-days of exposure data were collected for the following pollutants: PM2.5, 

PM10, O3, NO2, SO2, elemental carbon, organic carbon, and volatile organic compounds (Sarnat 

et al., 2001). A subset of PM2.5 filters was analyzed for SO4

2- concentration to estimate the 

personal exposure to PM2.5 of ambient origin. The mean ambient 24-h avg O3 concentration was 

approximately 36 ppb. Sarnat et al. (2001) found that ambient 24-h avg O3 concentrations and 

ambient 24-h avg PM2.5 levels, both measured at centrally located monitoring sites, were 

positively associated in the summer ($ = 0.84 [95% CI: 0.56, 1.12], Spearman r = 0.67 

[p < 0.05]) and negatively associated in the winter ($ = !0.67 [95% CI: !0.91, !0.43], 

Spearman r = !0.72 [p < 0.05]). An association also was observed between ambient O3 

concentrations and personal PM2.5 of ambient origin, with a mixed regression effect estimate of 

$ = 0.37 (95% CI: 0.25, 0.49) in the summer and ß = !0.36 (95% CI: !0.41, !0.31) in the 

winter. However, no significant relationship was found between 24-h avg personal O3 exposure 

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and personal exposure to PM2.5 of ambient origin in either season ($ = 0.22 [95% CI: !0.06, 

0.50] in the summer and $ = !0.18 [95% CI: !0.39, 0.03] in the winter). 

In a related study conducted in Boston, MA, 20 healthy senior citizens and 23 school 

children were monitored during six sessions for a total of 714 person-days in both the summer 

and winter seasons (Sarnat et al., 2005). Personal SO4

2- exposures were used as indicators of 

personal exposure to PM2.5 of ambient origin. Mean ambient 24-h avg O3 levels ranged from 

22.7 to 31.6 ppb during the three summer sessions and 14.0 to 21.8 ppb during the three winter 

sessions. Similar to the earlier study conducted in Baltimore, ambient 24-h avg O3 levels and 

ambient 24-h avg PM2.5 levels were positively associated in the summer ($ = 0.51 [95% CI: 

0.34, 0.68]) but negatively associated in the winter ($ = !0.53 [95% CI: !0.22, !0.85]). 

In addition, an association was observed between ambient O3 concentrations and personal PM2.5 

of ambient origin, but only in the summer ($ = 0.24 [95% CI: 0.20, 0.28]). However, in contrast 

to the Baltimore results, a significant association between 24-h avg personal O3 exposure and 

personal exposure to PM2.5 of ambient origin was observed both in the summer ($ = 0.35 [95% 

CI: 0.22, 0.47]) and in the winter ($ = 0.07 [95% CI: 0.01, 0.13]). Neither of the Sarnat et al. 

(2001, 2005) studies evaluated relationships between personal O3 exposures and personal 

exposure to other gaseous pollutants. 

Results from the two related Sarnat studies provide initial evidence that the relationship 

between personal O3 exposure and personal exposure to PM2.5 of ambient origin may differ by 

city/region and season. The authors noted that their results were for a small, nonrandom 

selection of subjects living in the eastern United States, therefore caution should be exercised in 

generalizing the results to other locations and cohorts. The results are likely to vary for locations 

depending on indoor air exchange rates and amount of time spent outdoors. 

7.6.4.2 Assessment of Confounding Using Multipollutant Regression Models 

Multipollutant regression models are generally used to determine whether the pollutantspecific effect is robust. However, because of the multicollinearity among O3 and pollutants and 

the changing correlations by season, multipollutant models may not adequately adjust for 

potential confounding, especially when using year-round data. These limitations need to be 

considered when evaluating results from multipollutant models. Results from the U.S. 90 cities 

study, which included 80 cities with O3 data, indicated that while the addition of PM10 in the 

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model did not substantially change the O3-mortality risk estimates, slight declines in the O3 

effect were observed, as shown in Figure 7-31 (Samet et al., 2000; reanalysis Dominici et al., 

2003). In the extended U.S. 95 communities study (Bell et al., 2004), the city-specific O3­

mortality effects were robust to adjustment for PM10, as indicated by the nearly 1:1 ratio between 

estimates with and without PM10 adjustment shown in Figure 7-32. This finding suggested 

that PM10 generally did not confound the association between O3 and mortality. Limited data 

were available to examine the potential confounding effect of PM2.5 on the O3-mortality 

relationship. A weighted second-stage linear regression indicated that there was no association 

between long-term PM2.5 average and the community-specific O3-mortality effect estimate. 

Several other mortality and morbidity studies have investigated confounding of O3 risk estimates 

using multipollutant models with year-round data, and most have reported that O3 effects were 

robust to adjustment for copollutants (see Figures 7-11 and 7-22 in Sections 7.3.3 and 7.4.6, 

respectively). 

Figure 7-31. Posterior means and 95% probability intervals for the national average 

estimate of O3 effects on total mortality from non-external causes per 10 ppb 

increase in 24-h avg O3 at 0-, 1-, and 2-day lags within sets of 80 U.S. cities 

with pollutant data available. Models A = O3 only; B = O3 + PM10; C = O3 + 

PM10 + NO2; D = O3 + PM10 + SO2; E = O3 + PM10 + CO. 

Source: Derived from Dominici et al. (2003). 

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Figure 7-32. Maximum likelihood estimates of O3-mortality for 95 U.S. communities, 

determined using a constrained distributed lag model for lags 0 through 

6 days. Same data set was used for O3 estimates with and without 

adjustment for PM10. 

Source: Derived from Bell et al. (2004). 

The pollutant most correlated with O3 in the summer is sulfate (which is in the fine particle 

size range), especially in the eastern United States. Therefore, the main potential confounders of 

interest for O3 in the summer are PM2.5 and sulfate. Once again, the results from two-pollutant 

regression models with O3 and sulfate (or PM2.5) should be interpreted with caution because both 

of these pollutants are formed under the same atmospheric conditions and both are part of the 

“summer haze” pollution mix. A simple two-pollutant regression model does not address their 

possible synergistic effects, and the high correlation between the two pollutants may lead to 

unstable and possibly misleading results. In any case, most studies that analyzed O3 with PM 

indices did not have PM2.5 data and very few examined sulfate data. A mortality study by Lipfert 

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et al. (2000a) found that all-year O3 risk estimates were not affected by the addition of sulfate 

(3.2% versus 3.0% with sulfate per 45 ppb less background level of 1-h max O3). Among 

studies with PM2.5 data, including Santa Clara County, CA (Fairley, 1999; reanalysis Fairley, 

2003), Philadelphia, PA (Lipfert et al., 2000a), and Detroit, MI (Lippmann et al., 2000; 

reanalysis Ito, 2003), most examined copollutant models for year-round data only, but 

O3-mortality risk estimates were not substantially affected by the addition of PM2.5. An analysis 

of Philadelphia and Detroit data by season suggested that O3-mortality risk estimates were not 

sensitive to adjustment for PM2.5 in all-year or seasonal analyses (Ito et al., 2005). Including 

both O3 and PM2.5 in the regression models for these cities tended to attenuate both pollutants’ 

coefficients, but not substantially. The models with both O3 and PM2.5 in these cities often 

showed better fits (i.e., lower Akaike’s Information Criteria) than those with either pollutant 

alone. Ito et al. (2003) concluded that these results suggest that O3 and PM2.5 contribute 

independently to mortality. 

Other studies have estimated O3 health risks with copollutants in the model by season. 

Respiratory hospitalization studies conducted during the warm season in Canada observed 

consistent O3 risk estimates with the inclusion of PM2.5 in the model (Burnett et al., 1997b, 

2001). In one of these studies (Burnett et al., 1997b), the effect of O3 also was adjusted for 

sulfate. With the addition of sulfate in the model, the risk estimate for O3 on respiratory 

hospitalizations remained relatively stable, from a 14.4% (95% CI: 8.7, 20.5) excess risk to a 

11.7% (95% CI: 5.6, 18.0) excess risk per 25 ppb increase in 12-h avg O3 at a 1- to 3-day lag. 

In contrast, the effects for sulfate were reduced in half after adjusting for O3. Amongst the 

mortality studies (see Figure 7-23 in Section 7.4.6), adjusting for copollutants, in particular PM 

indices, did not substantially change the warm-season O3-mortality effect estimates, with both 

slight reductions and increases observed in the adjusted estimates. In the analysis using coolseason-only data, the O3 effect estimates were generally negative, but none were statistically 

significant. The O3 risk estimates all increased slightly with the adjustment of PM indices. 

The inverse relationship between O3 and PM during the cool season most likely influenced 

the O3-mortality effect estimates in the single-pollutant models. 

In field studies, power to assess independent O3 effects may be limited by small sample 

sizes and short follow-up times. Yet, the O3 effect also was robust to the addition of copollutants 

in multipollutant models, with a few exceptions. For example, the effect of O3 on PEF was not 

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robust to adjustments for PM2.5 and sulfate, in studies by Romieu et al. (1996) and Neas et al. 

(1999). In general, however, O3 effects on respiratory symptoms (Romieu et al., 1996), lung 

function parameters (Brauer et al., 1996; Gold et al., 1999), and asthma medication use (Gent 

et al., 2003) were robust to inclusion of PM2.5. Further, the effects for O3 were observed to be 

stronger than those for PM. 

Multipollutant regression analyses indicated that O3 risk estimates, in general, were not 

sensitive to the inclusion of copollutants, including PM2.5 and sulfate. These results suggest that 

the effect of O3 on respiratory health outcomes appears to be robust and independent of the 

effects of other copollutants. 

7.6.5 Concentration-Response Function and Threshold 

An important consideration in characterizing the public health impacts associated with O3 

exposure is whether the concentration-response relationship is linear across the full 

concentration range or instead shows nonlinearity. Of particular interest is the shape of the 

concentration-response curve at and below the level of the current 8-h standard of 80 ppb. There 

are limitations to identifying possible “thresholds” in air pollution epidemiologic studies, 

including difficulties related to the low data density in the lower concentration range, possible 

influence of measurement error, and individual differences in susceptibility to air pollution 

health effects. Despite these difficulties, the slope of the O3 concentration-response relationship 

has been explored in several studies. 

To examine the shape of the concentration-response relationship between O3 and mortality, 

Gryparis et al. (2004) used meta-smoothing to combine smooth curves across the 23 European 

cities in a hierarchical model. The mean 1-h max O3 concentrations ranged from 44 ppb in 

Tel Aviv to 117 ppb in Torino, Italy during the summer. For the summer period, while the 

estimated concentration-response curve did not appear to deviate significantly from linearity, 

there were indications of decreasing effects at lower exposures. 

In the U.S. 95 communities study (Bell et al., 2004), effect estimates calculated using only 

days with 24-h avg O3 levels less than 60 ppb were compared to those using all data. At a lag of 

1 day, O3 was associated with an excess risk of 0.36% (95% PI: 0.12, 0.60) per 20 ppb increase 

in 24-h avg O3 using data from all days and only a slightly smaller risk of 0.30% (95% PI: 0.08, 

0.54) when data were limited to days less than 60 ppb. These results suggest that if there is a 

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threshold, it must be notably lower than a 24-h avg O3 of 60 ppb. In a more recent study by Bell 

et al. (2006; published online January 23, 2006), the shape of the concentration-response curve 

for the O3-mortality relationship was evaluated in 98 U.S. urban communities for the period 

1987 to 2000. The measure of exposure was the average of the same and previous days’ ambient 

24-h avg O3 concentration. Analytical methods included linear, subset, threshold, and spline 

models. Results from all methods indicated that if a threshold did exist, it would have to be at 

low concentrations, less than a 24-h avg O3 of 15 ppb. 

Fairley (2003) reanalyzed the Santa Clara County mortality data using GAM with stringent 

convergence criteria and examined a new exposure index for O3. He noted O3 concentrations 

exceeding 60 ppb each hour and calculated a daily sum of these exceedances. Fairley’s index 

incorporates measures of concentration and exposure duration; thus, this index represents a 

linear time-integrated concentration, also known as dosage. The O3 index with the 60 ppb 

“threshold level” was found to be significantly associated with mortality in single-pollutant 

models as well as in multipollutant models. Two other “threshold levels” were examined, 

40 ppb and 80 ppb. Both produced statistically significant results in single-pollutant models. 

These results suggest that the threshold for O3-mortality effects, if it exists, is likely less than 

40 ppb. The implication for thresholds in terms of the three standard indices (i.e., 1-h max, 8-h 

max, and 24-h avg) is unclear, but there may be an empirical relationship. 

Vedal et al. (2003) observed that the annual mean 1-h daily max O3 concentration of 

27.3 ppb (SD 10.2) in Vancouver, Canada, was lower than that in any of the 80 NMMAPS cities

(Samet et al., 2000); thus, a Vancouver study might provide a better focus on the shape of the O3 

concentration-response curve at lower levels. An O3 effect was observed on total mortality at a 

0-day lag during the summer. Ozone effects on respiratory mortality at a 2-day lag and 

cardiovascular mortality at a 0-day lag also were observed in the summer in this study. The 

effect of O3 on mortality was robust in two-pollutant models. Vedal et al. (2003) concluded 

that O3 concentrations were associated with adverse effects on mortality even at low levels. 

Although this study appears to support the argument that there are no threshold concentrations 

below which adverse effects cannot be detected, the results must be interpreted with caution as 

concerns remain. Vedal et al. (2003), for example, questioned if O3, other gaseous pollutants, 

and PM may be acting as surrogate markers of pollutant mixes that contain more toxic 

compounds, since the low measured concentrations were unlikely, in their opinion, to cause the 

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observed effects. They further stated that measurement error and interference by meteorological 

factors might have contributed to the inability to detect a threshold. 

Kim et al. (2004) investigated the presence of a threshold in O3-mortality effects in Seoul, 

Korea by analyzing data using a log linear GAM (linear model), a cubic natural spline model 

(nonlinear model), and a B-mode splined model (threshold model). Models were stratified by 

season and adjusted for PM10, long-term time trend, and meteorological variables. The mean 1-h 

max O3 levels were 46.9 ppb (SD 22.5) during the summer and 21.3 ppb (SD 6.9) during the 

winter. Threshold values were estimated as 1-h max O3 levels of 28 ppb when using all-year 

data and 45 ppb for summer-only data. None of the other pollutants examined, including PM10, 

SO2, NO2, and CO, had a nonlinear association with mortality. Using summer-only data, the 

B-spline model resulted in an excess mortality risk of 7.1% (95% CI: 3.1, 11.2) per 40 ppb 

increase in 1-h max O3, compared to an excess risk of 3.6% (95% CI, 0.5, 6.8) calculated using 

the log linear model. If a threshold truly exists, results from the Kim et al. (2004) study suggest 

that the use of log-linear models may underestimate the O3 effect on mortality at levels above the 

threshold. 

Other studies examining the effect of O3 on mortality also have found suggestive evidence 

for a possible threshold level. In a London, England study (Anderson et al., 1996), an adjusted 

O3-mortality bubble plot suggested that a threshold might exist around 50 ppb for 8-h avg O3. 

A study by Simpson et al. (1997) in Brisbane, Australia observed a significant excess risk in 

mortality only in the highest quintile of O3 exposure, which had a mean concentration of 42 ppb 

for 1-h max O3. 

Among several studies with morbidity outcomes, examination of the shape of the 

concentration-response function indicated evidence of an effect threshold. In a study of all-age 

respiratory hospital admissions in Toronto, Canada, effects of O3 appeared to become apparent 

only above an approximate 30 ppb daily 1-h max O3 (Burnett et al., 1997b). Also, Ponce de 

Leon et al. (1996) reported an indication of a threshold in the O3 effect on hospitalizations in 

London, England at 40 to 50 ppb for 8-h max O3 and 50 to 60 ppb for 1-h max O3. In a study of 

emergency department visits for asthma in St. John, Canada, effects observed in the over 

15 years age group were apparent only when data above the 95th percentile (75 ppb daily 1-h 

max O3) were included (Stieb et al., 1996). However, other morbidity studies observed a 

monotonic increase in the concentration-response function, suggesting that there was no 

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threshold in O3 effects on hospitalizations and emergency department visits (Burnett et al., 

1997a; Jaffe et al., 2003; Petroeschevsky et al., 2001; Tenías et al., 1998). 

In a field study by Mortimer et al. (2002), the associations of ambient O3 levels with PEF 

and asthma symptoms were investigated in eight urban cities in the United States. The mean 8-h 

avg O3 was 48 ppb, with less than 5% of days exceeding 80 ppb. Analysis performed using all 

data indicated that a 15 ppb change in 8-h avg O3 was associated with decrements in PEF 

(!0.59% [95% CI: !1.05, !0.13]) and increased incidence of respiratory symptoms (odds ratio 

of 1.16 [95% CI: 1.02, 1.30]) over multiday lag periods. When data were restricted to days 

when ambient O3 concentrations were less than 80 ppb, the O3 effects persisted, with a 

significant PEF decline (!0.70% [95% CI: !1.29, !0.12]) and incidence of morning symptoms 

(odds ratio of 1.17 [95% CI: 1.01, 1.35]). A study by Chen et al. (1999) also found that there 

was no clear threshold in the O3 effect on FEV1 and FVC in Taiwanese school children. The 1-h 

max O3 concentration ranged from 19.7 to 110.3 ppb. 

The studies of both Brauer et al. (1996) and Korrick et al. (1998) demonstrate that 

exposure duration and exercise level, in addition to O3 concentration, must be considered when 

evaluating thresholds. In the study by Brauer et al., the mean O3 concentration during the 

11-hour work shift was 26.0 ppb (SD 11.8). Workers experienced a change of !180.0 mL (95% 

CI: !227.0, !133.0) in FEV1 levels the next morning per 40 ppb increase in 1-h max O3. The 

hikers in the study by Korrick et al. (1998) were exposed to mean O3 levels of 40 ppb (SD 12) 

over the duration of their hike (mean 8 hours). Korrick et al. observed a mean change of 

!62.5 mL (95% CI: !115.3, !9.7) in pre-hike to post-hike FEV1 per 30 ppb increase in 8-h 

avg O3 when all hikers were included in the analysis; however, when analysis was restricted to 

hikers with wheeze or asthma, a larger change of !182.5 mL (95% CI: !312.2, !52.9) was 

observed. In both studies, large reductions in lung function were observed in subjects exposed to 

relatively low levels of O3 over multiple hours while active outdoors. 

Other studies that may provide information concerning the concentration-response 

relationship are those that reported larger excess health risks in the warm (or summer) season 

than in the cold (or winter) season (as discussed in Section 7.6.3.2). During the cold season, 

when O3 concentrations are generally low, no consistent O3 effects were observed across the 

various health outcomes. These results appear to provide epidemiologic evidence of the 

presence of a threshold in O3 health effects. However, other factors also may contribute to the 

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observed lack of an association between O3 and health outcomes in the cold season. First, 

potential confounding by copollutants may mask the effect of O3 during the cool season. For 

example, O3 levels have been found to be inversely related to PM concentrations during the 

winter in some cities, e.g., New York City and Philadelphia (Ito et al., 2005). If PM is also 

associated with the health outcome of interest, inadequate control of confounding of the PM 

effect would result in negative or null associations for O3 during the winter. In the few mortality 

studies that adjusted for PM in season-stratified analyses (see Figure 7-23), the cool-season O3 

risk estimates all increased slightly with the adjustment of PM indices; however, it should be 

noted that none achieved statistical significance. A second possible reason for the difference 

in O3 effects observed by season may be the changing relationship between O3 concentrations 

and personal exposure across seasons. The ambient O3 levels are lower in the cold season, but 

people are likely to be exposed to even lower levels of O3 during this season because of the 

shorter time spent outdoors and the longer time spent indoors with closed windows. Sarnat et al. 

(2005) observed that a 1 ppb increase in ambient O3 concentration was associated with a 

0.27 ppb (95% CI: 0.18, 0.37) increase in person O3 exposure during the summer and 0.04 ppb 

(95% CI: 0.00, 0.07) increase during the winter. 

A more “representative” concentration-response relationship may need to be examined in a 

summer-only data set as personal exposures tend to be better represented by ambient O3 

concentrations during that season. Even for summer data, however, an interpretation of the 

relationship is not straightforward because of the possible influence of the use of air conditioning 

(an effective remover of O3). Greater use of air conditioning is expected on hot days when 

the O3 level is higher, but the use of air conditioning may also vary from city to city and across 

social class within a city. Brauer et al. (2002) observed that surrogate measures of exposure (i.e., 

those from centrally-located ambient monitors) that were not highly correlated with personal 

exposures obscured the presence of thresholds in epidemiologic studies at the population level, 

even if a common threshold exists for individuals within the population. 

Obscuring of thresholds would be even greater if thresholds varied across individuals 

(Brauer et al., 2002). Effects occur at the molecular level and are eventually manifested at the 

clinical level. One possible molecular mechanism for adverse physiologic effects from O3 is its 

reaction with the carbon double bonds of lipids in the lung lining fluid (Levy et al., 2001). The 

molecular effects of O3 are partially mitigated by the antioxidant defense system, which varies in 

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individuals with age, diet, genetic factors, and disease status. This raises the possibility that 

there may be a level below which O3 would have few or no adverse effects as well as 

contributing to individual variability in the threshold level. The antioxidant defense system is 

one of many factors that contribute to the large intersubject variability in lung function responses 

to O3 reported in human clinical studies (discussed in Sections 6.4 and 6.5). As individual 

sensitivities to O3 health effect vary, a common threshold may not be observable at the 

population level in epidemiologic studies. 

As noted earlier, demonstrating a clear threshold at a population level is difficult due to 

low data density in the lower O3 concentration range, measurement error resulting from personto-person differences in the relationship between personal exposures and monitored ambient 

concentrations, and individual differences in susceptibility to air pollution health effects. From 

1990 to 2004, the 10th percentile values (which represent the lower concentration range) of the 

nationwide mean daily 8-h max O3 concentrations were approximately 40 ppb during the warm 

season (May to September) (see Figure 3-17 in Section 3.2). While no fully confident 

conclusion can be made regarding the threshold issue from epidemiologic studies alone, the 

limited currently available evidence suggests that if a population threshold level exists in O3 

health effects, it is likely near the lower limit of ambient O3 concentrations in the United States. 

7.6.6 Heterogeneity of Ozone Health Effects 

As described in Chapter 3 of this AQCD, O3 concentrations tend to be more spatially 

variable than PM2.5 concentrations in urban areas. In addition, relative personal exposures to O3 

likely vary by region. The geographic variability in O3 concentrations and personal exposures 

may contribute to the heterogeneity in observed O3 health effects. The degree of influence of the 

geographic variability on heterogeneity in effects will vary by study, as study design affects 

different aspects of exposure (e.g., time period and duration of exposure). 

More than 80% of the O3-mortality estimates from the various studies conducted in North 

America, South America, Europe, and Australia were between 0.5 and 5% excess risk per 40 ppb 

increase in 1-h max O3 using year-round data. In general, the O3-mortality estimates were 

greater when using summer only data compared to year-round data. Though not all statistically 

significant, most of the O3-mortality estimates were greater than zero, indicating a positive 

relationship between O3 exposure and mortality. The O3 risk estimates from the numerous 

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hospitalization and emergency department visit studies were generally larger in magnitude and 

more variable from study to study compared to the mortality studies. These differences in 

the O3 effect estimates may be attributable to the greater variability in the outcome measure in 

hospitalization studies compared to mortality studies, as a result of the use of more subcategories 

of outcome and varying degrees of severity. 

Three recent meta-analyses that included both U.S. and non-U.S. studies found consistent 

all-year combined point estimates: 1.75% (95% PI: 1.10, 2.40), 1.6% (95% CI: 1.1, 2.0), and 

1.64% (95% CI: 1.25, 2.03) per 20 ppb increase in 24-h average O3, for Bell et al. (2005), 

Ito et al. (2005), and Levy et al. (2005), respectively. Bell et al. further observed that the pooled 

estimate for U.S. studies (11 estimates), 1.69% (95% PI: 0.94, 2.78), was similar to the pooled 

estimate for the non-U.S. studies (30 estimates), 1.85% (95% PI: 0.94, 2.78). Levy et al. 

compared North American studies to European studies and also found nearly identical effect 

estimates. 

As differences in study design, population, and data analysis may affect risk estimates, 

studies that were conducted in multiple cities using standardized methods were further examined 

to investigate the geographic heterogeneity of O3 effects. Bell et al. (2004) conducted a timeseries analysis of O3 and mortality in 95 U.S. communities from 1987 to 2000. A 20-ppb 

increase in 24-h avg O3 levels in the previous week was associated with an increase of 1.04% 

(95% PI: 0.54, 1.55) excess risk of mortality in the pooled analysis of 95 communities using all 

available data. The median 24-h avg O3 concentrations varied from 14.4 ppb in Newark, NJ to 

37.3 ppb in Bakersfield, CA. Intercommunity heterogeneity in O3 effects was observed among 

the 95 communities (see Figure 7-17 of Section 7.4.3), which the authors noted as plausible 

given the community-specific differences in pollution characteristics, the use of air conditioning, 

time-activity patterns, and socioeconomic factors. One factor that appears to explain some of the 

intercity heterogeneity in the 95 U.S. communities study is long-term O3 concentration. 

A weighted regression of the community-specific mortality risk from acute O3 exposure and 

long-term average 24-h avg O3 was significant (p = 0.01). These results indicated a higher 

excess risk of mortality per incremental change in 24-h avg O3 in the previous week was 

observed in communities with lower long-term average O3. Note that this analysis only includes 

the 40 communities with warm-season data. This relationship can be explained in a three 

parameter Gompertz concentration-response model of the cumulative percent excess risk in 

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mortality by long-term average 24-h avg O3 levels (Figure 7-33). The method of estimating this 

curve is discussed in Annex Section AX7.3 along with limitations to inference concerning a 

threshold. The curve has shallower slopes above the median effective concentration (inflection 

point) than below that concentration point. The horizontal line is the 95% confidence interval on 

the long-term average 24-h avg O3 associated with a 0.05% excess risk in mortality (EC0.05%). 

The EC0.05% is a risk that is believed to be low enough to be in the noise band. The Gompertz 

model presents a threshold of effects at low concentrations and a flat curve at higher 

concentrations. Thus, to the extent that the estimated concentration associated with a 0.05% 

excess risk is realistic, the fit of the model supports the observed relationship between greater 

mortality risk per incremental change in acute O3 exposure at lower long-term O3 levels. 

Figure 7-33. The fitted Gompertz model of the cumulative percent excess risk in mortality 

by long-term average 24-h avg O3 concentrations using data from the 95 U.S. 

communities study (Bell et al., 2004). Only the 40 communities with warmseason data are included. Highlighted is the actual data range of the longterm average 24-h avg O3 concentrations. The horizontal line is the 95% CI 

on the long-term average 24-h avg O3 associated with a 0.05% excess risk in 

mortality (EC0.05%). 

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Greater heterogeneity was observed in the European study of 23 cities in 14 countries 

(Gryparis et al., 2004). In the year-round analyses, only 8 of the 23 cities had positive 

O3-mortality effect estimates. However, in the analyses using summer-only data, the risk 

estimates were positive in 19 of the 23 cities, with a range of 0.8 to 8% excess risk per 40 ppb 

increase in 1-h max O3. Median 1-h max O3 concentrations ranged from 44 ppb in Tel Aviv, 

Israel to 117 ppb in Torino, Italy during the summer. The heterogeneity may be attributable to 

the considerable variability among countries in factors that may influence the relationship 

between ambient O3 concentrations and personal exposure to O3, such as climate, use of air 

conditioning, personal activity patterns, and socioeconomic factors. In addition, the variability 

in the concentration and composition of copollutants by cities or countries may contribute to the 

heterogeneity in the O3-mortality effects. For example, concentrations of NO2 may vary widely 

by region, depending on the differences in traffic density. 

Among the hospitalization studies, Burnett et al. (1997a) conducted the largest study of 

16 Canadian cities. The mean daily 1-h max O3 was 31 ppb (range 26-38) in the 16 cities. The 

pooled O3 estimate was 5.6% (95% CI: 3.4, 7.9) excess risk in respiratory hospitalization per 

40 ppb increase in 1-h max O3 using warm-season data (April to December). The risk estimates 

were fairly homogenous across the 16 Canadian cities, ranging from 3.1% for Vancouver to 

7.7% for Quebec City. 

Anderson et al. (1997) investigated the association between O3 and hospital admissions for 

COPD in five European cities (London, Paris, Amsterdam, Rotterdam, and Barcelona). The 

pooled effect estimate was 5.0% (95% CI: 2.6, 7.6) excess risk per 30 ppb increase in 8-h 

max O3 for year-round data. Results from the APHEA study showed similar variability to that 

from the Burnett et al. (1997a) study. The year-round effects estimates were lower in the two 

Dutch cities (2.5% excess risk) compared to that in Paris (7.7% excess risk); however, analyses 

indicated that there was no significant heterogeneity in effects by city. The authors further noted 

that among the pollutants examined (O3, BS, TSP, SO2, and NO2), O3 had the most consistent 

and significant findings. 

Among the field studies, various respiratory health outcomes were examined, including 

PEF, spirometric parameters, respiratory symptoms, and medication use. Only one field study 

investigated the O3 effect in several locations (Mortimer et al., 2002). Mortimer et al. (2002) 

investigated the association of ambient O3 concentrations with PEF and asthma symptoms in 

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asthmatic children living in eight urban cities in the United States: St. Louis, MO; Chicago, IL; 

Detroit, MI; Cleveland, OH; Washington, DC; Baltimore, MD; East Harlem, NY; and Bronx, 

NY. In the analysis pooling data from all eight cities, a 30 ppb increase in 8-h avg O3 was 

associated with a decrement of !1.18% (95%CI: !2.10, !0.26) in morning PEF for a 5-day 

cumulative lag period. The percent changes in PEF were negative in all cities except for 

Baltimore, 0.49%. Among the other seven cities, the percent changes in PEF were quite 

homogenous, with values ranging from !1.08% for Washington, DC to !1.71% for St. Louis. 

A 30 ppb increase in 8-h avg O3 also was associated with an increased incidence of morning 

symptoms in the pooled analysis (odds ratio of 1.35 [95% CI: 1.04, 1.69] for a 4-day cumulative 

lag period). In all cities except for St. Louis, there was an increase in the incidence of morning 

symptoms. In these cities, the odds ratios for incidence of morning symptoms were somewhat 

more varied than for the PEF measurements, ranging from 1.19 for Chicago to 2.96 for Detroit. 

Most of the multicity and meta-analyses studies consistently found positive associations 

between O3 and mortality. Consistent O3 effects on hospitalizations and various respiratory 

health outcomes also were found. The observed heterogeneity of O3 effects may be partially 

attributable to the use of centrally-located ambient monitors to assess exposure. There may be 

differences in relative personal exposures to O3 due to varying factors, such as use of air 

conditioning and activity patterns, that affect the relationship between personal exposure and 

ambient concentrations. For example, Levy et al. (2005) found suggestive evidence that air 

conditioning prevalence was a predictor of heterogeneity in O3 risk estimates in their metaanalysis. The variability in the concentration and composition of other pollutants present also 

may contribute to the heterogeneity of the effect of O3 on health outcomes as confounding by 

copollutants may vary by region. 

7.6.7 Health Effects of Ozone in Susceptible and Vulnerable Populations 

In this section, the effects of O3 on morbidity and mortality in potentially susceptible and 

vulnerable populations will be examined. In epidemiologic studies of O3 health effects, the most 

widely studied subpopulation was asthmatics. Also of interest were the observed health effects 

of O3 on different age groups, particularly children and the elderly. Other groups that are 

vulnerable to O3 health effects are those that spend a lot of time outdoors at higher exertion 

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levels, such as outdoor workers. This section begins with a discussion of the O3-related health 

effects in asthmatics. 

7.6.7.1 Health Effects Associated with Ambient Ozone Exposure in Asthmatics 

Epidemiologic studies of health effects from acute O3 exposure in asthmatics have 

examined a range of outcomes: pulmonary function, respiratory symptoms, inflammation, 

emergency room visits, hospital admissions, and mortality. Chronic O3 exposure studies have 

investigated similar outcomes, with the exception of emergency room visits and hospitalizations. 

Both are discussed in the earlier text. This subsection draws together this information to 

examine whether the evidence indicates that O3 exposure impacts asthmatics. 

In Germany and Mexico City, O3 exposure was associated with a decline in FEV1 in 

asthmatic adults and children (Höppe et al., 1995a, 2003; Romieu et al., 2002). Change in FEV1 

also was examined in a group of asthmatic hikers in Mount Washington, NH (Korrick et al., 

1998). The mean hourly O3 concentration during each hike was 40 ppb (range 21-74). 

Compared to the healthy subjects, the asthmatic subjects experienced a 4-fold greater decline in 

FEV1 with the same exposure to O3 (mean change of !1.08% [95% CI: !2.49, 0.33] versus 

!4.47% [95% CI: !7.65, !1.29] per 30 ppb increase in 8-h avg O3). The results from the hiker 

study are consistent with those observed in a controlled human exposure study, which observed 

an approximately 2-fold greater decrement in FEV1 among mild-to-moderate asthmatics versus 

nonasthmatic subjects performing light exercise during a 7.6-h exposure period at 0.16 ppm O3 

(Horstman et al., 1995). 

PEF was examined in panels of asthmatic children in several field studies (see Figures 7-1 

and 7-2). Collectively, most of the studies indicated decrements of morning PEF, though only a 

few estimates were statistically significant. One multicity study of eight urban areas in the 

United States observed O3-related reductions in morning PEF that were not significant in each 

individual city (Mortimer et al., 2002); however, the analysis combining data from all eight cities 

indicated a significant decline in PEF with a cumulative lag of 1 to 5 days of O3 exposure. The 

median 8-h avg O3 levels ranged from 34 to 58 ppb across the eight cities. The odds ratio for the 

incidence of $10% decline in morning PEF was greater than one, which was discussed by the 

author as an indication that O3 exposure might be associated with clinically important changes in 

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PEF in asthmatic children. The study examined 846 asthmatic children, the largest asthma panel 

study reported. 

Mortimer et al. (2000) observed that the subpopulation of asthmatic children with a history 

of low birth weight or premature birth had greater O3-associated declines in PEF (mean change 

of !3.66% [95% CI: !5.30, !2.02] per 30 ppb increase in 8-h avg O3) than normal birth weight 

children (!0.60% [95% CI: !1.58, 0.38]). Low birth weight and prematurity are associated with 

reduced lung function, higher levels of airway reactivity and increased susceptibility to lung 

damage (Barker et al., 1993; Rona et al., 1993), which may explain why these factors are found 

to increase susceptibility to respiratory insults of air pollution in children. 

Lung function parameters have been evaluated for clinical significance. A reversible 5 to 

15% decline in FEV1 in an individual may have clinical importance to asthma morbidity 

(American Thoracic Society, 1991; Lebowitz et al., 1987; Lippmann, 1988). The National 

Institutes of Health (1997) has stated that PEF below 80% of the personal best indicates a need 

for additional medication use in asthmatics. At a population level the mean changes in lung 

function attributable to O3 exposure do not generally exceed 10% changes in FEV1 or PEF per 

standardized increment of O3. At an individual level, a subpopulation of susceptible asthmatics 

are likely experiencing clinically significant declines in lung function. Höppe et al. (2003) 

investigated the effects of O3 on the lung function of potential risk groups, including asthmatics, 

children, athletes, and the elderly. The mean afternoon (1 p.m.-4 p.m.) 1⁄2-h max O3 levels on 

high O3 days were 66.9 ppb (range 51-91) for the asthmatics and 65.2 ppb (range 43-88) for the 

children. The afternoon 1⁄2-h max O3 levels in these two groups were similar to those 

experienced by the elderly and athletes. Consistent associations between O3 and group means of 

lung function endpoints were not observed for the various risk groups. However, a potential 

pattern of O3 sensitivity was observed when individual data was examined. About 20% of the 

asthmatics and children were regarded as O3 responders (i.e., individuals with >10% change 

in FEV1) compared to only 5% of the elderly and athletes. In these responders, a significant O3 

concentration-response relationship was observed in the regressions using repeated 

measurements from individuals. These results indicated that while the population as a whole 

was not reacting to O3, susceptible individuals were experiencing clinically significant declines 

in lung function in response to O3 exposure. 

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Respiratory symptom increases in asthma panels were examined in several field studies, 

some of which also examined PEF as discussed above. The outcome definition of symptoms 

varied among these studies. Collectively, the results are suggestive of a potential O3 effect on 

respiratory symptoms, but the evidence is not strong in the available studies. Two U.S. studies 

that examined larger panels might be better to draw inferences from as the large sample size 

provided greater power to examine the effect of O3 on respiratory symptoms. The eight U.S. 

urban cities study reported that morning symptoms in the 846 asthmatic children were most 

strongly associated with a 4-day cumulative lag of O3 concentrations (Mortimer et al., 2002). 

A New England study examined 271 asthmatic children and observed an O3 effect on a variety 

of respiratory symptoms at a lag of 1 day among the 130 subjects who used maintenance asthma 

medications (Gent et al., 2003). The mean 1-h max O3 was 58.6 ppb (SD 19.0). 

Only a few epidemiologic studies have examined airway inflammation in asthmatics. 

A Mexico City study indicated that supplementation with antioxidants may modulate the impact 

of O3 exposure on the small airways of children with moderate-to-severe asthma (Romieu et al., 

2002). The mean 1-h max O3 was 102 ppb (SD 47). A related study indicated that asthmatic 

children with GSTM1 null genotype were found to be more susceptible to the impact of O3 

exposure on small airways (Romieu et al., 2004). 

Emergency department visits for asthmatics have been examined in several studies and 

range from negative to positive results (see Figure 7-8 in Section 7.3.2). Studies of mostly 

year-long data tended to produce inconsistent results, with some finding negative estimates 

(Atkinson et al., 1999a; Castellsague et al., 1995; Thompson et al., 2001; Tobías et al., 1999). 

Warm-season studies tended to yield positive outcomes, as expected based on earlier 

discussions. Two studies in Atlanta, GA (Tolbert et al., 2000) and Valencia, Spain (Tenías et al., 

1998) indicated positive effects in warm-season analyses. Further, a Canadian study, one of the 

larger studies conducted in the summertime, reported a large increase in asthma emergency 

department visits when the daily 1-h max O3 concentration exceeded 75 ppb (Stieb et al., 1996). 

A three-city study in Ohio also indicated an increased risk of asthma visits during the summer 

(Jaffe et al., 2003). The mean 8-h max O3 levels ranged from 50 to 60 ppb in the three cities. 

Hospital admission studies that specifically examined asthmatics were fewer in number 

than those that examined total respiratory diseases. Associations were noted in all age groups in 

studies conducted in Seattle, WA (Sheppard, 2003), New Jersey (Weisel et al., 2002), Toronto, 

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Canada (Burnett et al., 1999), London, England (Anderson et al., 1998), Brisbane, Australia 

(Petroeschevsky et al., 2001), and Hong Kong (Wong et al., 1999a). However, several other 

studies, mostly examining the effect of O3 on asthmatic children, did not observe a significant 

relationship (Gouveia and Fletcher, 2000a; Lin et al., 2003; Morgan et al., 1998; Nauenberg and 

Basu, 1999; Schouten et al., 1996). 

Acute mortality related to asthma was examined in Barcelona, Spain (Saez et al., 1999; 

Sunyer et al., 2002). In the study by Sunyer et al. (2002), severe asthmatics with more than one 

asthma emergency visit were found to have the strongest mortality associations with O3. The 

median 1-h max O3 level was 35.8 ppb (range 3.4-146.1). 

Recent reports from longitudinal cohort studies in California have reported associations 

between the onset of asthma and long-term O3 exposures (Greer et al., 1993; McConnell et al., 

2002; McDonnell et al., 1999). In adult studies, associations were seen in males but not females 

(Greer et al., 1993; McDonnell et al., 1999). Among children residing in high O3 communities 

(mean 8-h avg O3 of 59.6 ppb [range 55.8-69.0]), McConnell et al. (2002) observed that asthma 

risk was elevated for those who played three or more sports as compared with those who did not 

play sports. Playing sports may indicate outdoor activity and an increased ventilation rate which 

leads to increased dose of O3. These outcomes would benefit from replication in other cohorts in 

regards to indicating weight of a causal interpretation. 

A few studies provide limited discussion of concentration-response functions and 

thresholds. In the eight U.S. urban cities study of asthmatic children, the odds ratios for 

incidence of $10% decline in morning PEF and incidence of morning symptoms when excluding 

days with 8-h avg O3 greater than 80 ppb were nearly identical to those including data from all 

days (Mortimer et al., 2002). In the New England asthma panel study (Gent et al., 2003), some 

of the associations for symptoms occurred at 1-h max O3 levels below 60 ppb. In the St. John, 

Canada study (Stieb et al., 2003), an effect of O3 on emergency department visits was reported 

with evidence of a threshold somewhere in the range below a 1-h max O3 of 75 ppb in the 

15 years and over age group. 

Overall, subjects with asthma have been examined across most health endpoints of interest. 

The results reported in these studies vary with some indicating a positive excess risk associated 

with O3. While no endpoint in itself seems to indicate an unquestionable demonstration of an 

association, studies with adequate sample size and power consistently provide positive results, 

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especially during the summer months when higher O3 levels occur. This view is strengthened as 

positive results are obtained cohesively across the varied outcomes. Therefore, based on the 

evidence, it seems prudent to consider asthmatics as a susceptible group that requires more 

protection from O3 exposures than the general public. 

7.6.7.2 Age-Related Differences in Ozone Effects 

The American Academy of Pediatrics (2004) notes that children and infants are among the 

most susceptible to many air pollutants, including O3. Eighty percent of alveoli are formed 

postnatally and changes in the lung continue through adolescence; the developing lung is highly 

susceptible to damage from exposure to environmental toxicants (Dietert et al., 2000). Children 

also have increased vulnerability as they spend more time outdoors, are highly active, and have a 

high minute ventilation, which collectively increase their dose (Plunkett et al., 1992; Wiley et al., 

1991a,b). In addition to children, the elderly are frequently classified as being particularly 

susceptible to air pollution. The basis of the increased sensitivity in the elderly is not known but 

one hypothesis is that it may be related to changes in the respiratory tract lining fluid antioxidant 

defense network (Kelly et al., 2003). 

Several mortality studies have investigated age-related differences in O3 effects. Among 

the studies that observed positive associations between O3 and mortality, a comparison of all age 

or younger age (#65 years of age) O3-mortality risk estimates to that of the elderly population 

(>65 years) indicates that, in general, the elderly population is more susceptible to O3 effects 

(Borja-Aburto et al. 1997; Bremner et al., 1999; Gouveia and Fletcher 2000b; O’Neill et al., 

2004; Simpson et al., 1997; Sartor et al., 1995; Sunyer et al., 2002). For example, a study by 

Gouveia and Fletcher (2000b) examined the O3-mortality effect by age in São Paulo, Brazil. The 

mean 1-h max O3 level was 35.1 ppb (SD 21.7). There were 151,756 deaths for all non-violent 

causes over the period of 1991 to 1993, of which 49% occurred in the elderly. Among all 

ages, O3 was associated with a 0.6% (95% CI: !0.8, 2.0) excess risk in all cause mortality per 

40 ppb increase in 1-h max O3. In comparison, in the elderly population, the O3-mortality risk 

estimate was nearly 3-fold greater, 1.7% (95% CI: 0.0, 3.3). Similarly, a Mexico City study 

found that O3-mortality risk estimates were 1.3% (95% CI: 0.04, 2.6) and 2.8% (95% CI: 1.0, 

4.6) per 20 ppb increase in 24-h avg O3 concentration in all ages and the elderly, respectively 

(O’Neill et al., 2004). The mean 24-h avg O3 level was 35.3 ppb (SD 11.0). 

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The meta-analysis by Bell et al. (2005) found a larger effect estimate for the elderly (2.92% 

[95% PI: 1.34, 4.51] per 20 ppb increase in 24-h avg O3) than for all ages (1.75% [95% PI: 

1.10, 2.37]). In the large U.S. 95 communities study (Bell et al., 2004), effect estimates were 

slightly higher for those aged 65 to 74 years, 1.40% (95% PI: 0.56, 2.25) excess risk per 20 ppb 

increase in 24-h avg O3, compared to individuals less than 65 years and 75 years or greater, 

1.00% (95% PI: 0.20, 1.85) and 1.04% (95% PI: 0.36, 1.75), respectively, using a constrained 

distributed 7-day lag model. Bell et al. (2004) notes that despite somewhat similar effect 

estimates, the absolute effect of O3 is substantially greater in the elderly population due to the 

higher underlying mortality rates, which leads to a larger number of extra deaths for the elderly 

compared to the general population. 

A few mortality studies examined another potentially susceptible age group, young 

children under the age of 5 years. The results were mixed, with one Mexico City study showing 

a lower risk of O3-related all cause mortality in young children compared to all ages and the 

elderly (Borja-Aburto et al., 1997) and one São Paulo, Brazil study showing a greater risk in 

respiratory mortality in young children compared to the elderly (Gouveia and Fletcher, 2000b). 

Another study in Mexico City by Loomis et al. (1999) observed a positive but nonsignificant 

association between ambient O3 concentrations and infant mortality (used Poisson GAM with 

default convergence criteria). Only a limited number of studies have focused on air pollution 

effects on mortality in children using a time-series approach. This is probably because the 

numbers involved are usually not adequate for such an analysis. Approximately 10% of 

mortality occurs in young children. Gouveia and Fletcher (2000b) noted that in the case of 

São Paulo, the mean number of daily deaths for respiratory causes in children was 2.2, which 

was unlikely to provide statistical power to detect the effects of air pollution even if they existed. 

In addition, there are other competing causes of mortality in young children, especially those in 

developing countries, which are together more important than air pollution. 

With respect to age-specificity of associations between O3 and acute respiratory 

hospitalizations or emergency department visits, no clear pattern emerges from recent studies. 

Associations have been reported for all ages (Anderson et al., 1997; Burnett et al., 1995, 1997b, 

1999; Weisel et al., 2002), adults or elderly (Burnett et al., 1997a; Delfino et al., 1997b, 1998b; 

Moolgavkar et al., 1997; Schwartz et al., 1996; Yang et al., 2003), and children (Burnett et al., 

2001; Gouveia and Fletcher, 2000a; Lin et al., 1999; Pönkä and Virtanen, 1996; Tolbert et al., 

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2000; Yang et al., 2003). Interestingly, studies that have examined effects in multiple age strata 

often have seen effects only in non-pediatric strata (Delfino et al., 1997b, 1998b; Stieb et al., 

1996; Jones et al., 1995). Several studies that focused on children did not report significant O3 

effects, though in some cases these studies are limited by small size, inadequate control of 

seasonal patterns, or very low O3 levels (Lierl and Hornung, 2003; Lin et al., 2003; Thompson 

et al., 2001). If O3 is causally related to exacerbations of respiratory diseases leading to hospital 

usage, one would expect to see effects most prominently among children, for whom asthma is 

more prevalent and O3 exposures may be greater. However, once again, other competing causes 

and the small numbers of hospitalizations in children likely limit the ability to examine O3­

related health effects. 

A few field studies compared the effect of O3 in different age groups. Korrick et al. (1998) 

examined changes in FEV1 and FVC related to O3 exposure in a group of hikers ranging in age 

from 18 to 64 years, and found that there was no association between O3 responsiveness and age. 

Brauer et al. (1996), in a study of berry pickers aged 10 to 69 years, also observed that subject 

age was not significantly associated with O3-related changes in lung function. However, a study 

by Höppe et al. (2003) observed that children, but not seniors (69 to 95 years of age), 

experienced a decline in lung function associated with O3 exposure. Approximately 20% of the 

children and juvenile asthmatics experienced a greater than 10% change in FEV1, compared to 

only 5% of the elderly population. The results by Höppe et al. are consistent with the 

diminishing responses to O3 exposure with increasing age observed in clinical studies. The 

clinical studies by Drechsler-Parks (1995) and Bedi et al. (1989) found that subjects aged 56 to 

89 years had markedly reduced responses to O3 exposure compared to young adults. 

Many field studies focused on the effect of O3 on the respiratory health of school children. 

In general, children experienced decrements in pulmonary function parameters, including 

PEF, FEV1, and FVC (Castillejos et al., 1995; Chen et al., 1999; Gielen et al., 1997; Gold et al., 

1999; Jalaludin et al., 2000; Mortimer et al., 2002; Romieu et al., 1996; Thurston et al., 1997). 

Increases in respiratory symptoms (Delfino et al., 2003; Gold et al., 1999; Neas et al., 1995; 

Romieu et al., 1996, 1997; Thurston et al., 1997) and asthma medication use (Delfino et al., 

1996; Just et al., 2002; Ostro et al., 2001; Thuston et al., 1997) also were observed in children. 

These respiratory health effects were largely observed in asthmatic children. Ozone-associated 

lung function declines were found in healthy children as well. 

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Collectively, there is supporting evidence of age-related differences in susceptibility 

to O3 health effects. The elderly population (>65 years of age) appear to be at increased risk 

of O3-related mortality and hospitalizations, and children (<18 years of age) experience other 

potentially adverse respiratory health outcomes with increased O3 exposure. One epidemiologic 

study also found that the lung function response to O3 exposure may be diminished in elderly 

populations; this finding is further supported by evidence from clinical studies. 

7.6.7.3 Vulnerability of Outdoor Workers and Others Who Participate in Outdoor 

Activities to Ozone Health Effects 

The health effects of O3 on outdoor workers and others who participate in outdoor 

activities have been investigated in various field studies. The most common endpoint examined 

in these studies is lung function. These individuals are typically exposed to high doses of O3 as 

they spend long hours outdoors often at elevated exertion levels. Their increased vulnerability 

to O3 health effects has been noted in epidemiologic studies. 

Brauer et al. (1996) repeatedly measured spirometric lung function before and after 

outdoor summer work shifts over 59 days on a group of 58 berry pickers (mean age 44 years 

[range 10-69]) in Fraser Valley, British Columbia, Canada. Outdoor work shifts averaged 

11 hours in duration. The mean ambient 1-h max O3 was 40.3 ppb (SD 15.2) over the study 

period. Heart rates during the work shift averaged 36% higher than resting levels, indicating 

elevated exertion levels while working outdoors. The lung function changes experienced by the 

workers in this study are large compared to those from other field studies (see Table 7-1b). The 

next morning FEV1 declined by 6.36% (95% CI: 4.70, 8.02) per 40 ppb increase in 1-h max O3. 

These results indicate that extended exposures to O3 at elevated exertion levels produce more 

marked effects on lung function. 

Mexico City outdoor street workers (n = 47) were repeatedly monitored for lung function 

changes at the end of the work shift over a two-month period (Romieu et al., 1998). Workers 

were exposed to outdoor ambient O3 levels for a mean of 7.4 hours during the day. The mean 

1-h max O3 was 123 ppb (SD 40). Among those who had never taken an antioxidant 

supplement, same day O3 concentrations were associated with decreases in afternoon FEV1. 

A mean change of !3.55% (95% CI: !6.28, !0.82) was observed per 40 ppb increase in 

1-h max O3. 

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Höppe et al. (1995a) examined forestry workers (n = 41) for O3-related changes in 

pulmonary function in Munich, Germany. The average time they spent outdoors was not 

presented. Ventilation rates were estimated from the average activity levels. Their ventilation 

rates were elevated at 40 L/min (compared to 15 L/min in the control group). When 

comparisons were made between high O3 days (mean 1⁄2-h max O3 of 64 ppb) and low O3 days 

(mean 1⁄2-h max O3 of 32 ppb), 59% of the forestry workers experienced a remarkable decrement 

in lung function (i.e., at least a 20% increase in specific airway resistance or at least a 10% 

decrease in FEV1, FVC, or PEF) on high O3 days. None experienced an improvement in lung 

function. A change of !56.0 mL (95% CI: !118.4, 6.4) in FEV1 was observed per 40 ppb 

increase in 1⁄2-h max O3. 

In addition to forestry workers, Höppe et al. (1995a) also monitored athletes (n = 43) in the 

afternoon following a two-hour outdoor training period. Athletes had a fairly high ventilation 

rate of 80 L/min. Compared to the forestry workers a smaller percentage of athletes experienced 

a remarkable decrement in lung function, 14%, on high O3 days; 19% of the athletes actually 

showed an improvement. Overall, a significant change in FEV1 was observed, !60.8 mL (95% 

CI: !115.2, !6.4) per 40 ppb increase in 1⁄2-h max, in the athletes. In a subsequent study, Höppe 

et al. (2003) reanalyzed the results of the athletes after stratifying the spirometric data by time of 

day (morning versus afternoon) and at different lag periods (lags of 0 to 2 days). The reanalysis 

indicated that O3-related decrements were observed only with the afternoon FEV1 at a 0-day lag, 

!1.26% (95% CI: !2.63, 0.10) change in FEV1 per 40 ppb increase in 1⁄2-h max O3. 

A study by Korrick et al. (1998) also examined the effects of multihour O3 exposures on 

adults exercising outdoors. A total of 530 hikers (mean age 35 years [range 18-64]) of Mount 

Washington, NH performed spirometry before and after hiking for a mean of 8 hours (range 

2–12). The mean of the hourly O3 concentrations during the hike was 40 ppb (range 21-74). 

After the hike, all subjects combined experienced a small mean decline of 1.53% (95% CI: 0.24, 

2.82) in FEV1 per 30 ppb increase in the mean of the hourly O3 concentrations during the hike. 

Ozone-related changes in lung function parameters were estimated after stratifying by hiking 

duration. Subjects who hiked 8 to 12 hours (n = 265) experienced a 2.07% (95% CI: 0.36, 3.78) 

decline in FEV1 per 30 ppb increase in the mean of the hourly O3 concentrations; those who 

hiked 2 to 8 hours (n = 265) experienced a smaller decline of 0.99% (95% CI: !0.72, 2.70). 

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Each hour hiked, which may reflect dose, was associated with a decline of 0.3% (p = 0.05) 

in FEV1, after adjusting for O3. 

The O3 effect attributable to exercise in children (n = 40) was investigated in a Mexico City 

study (Castillejos et al., 1995). These children were chronically exposed to moderate to high 

levels of O3. The mean ambient 1-h max O3 level during the study period was 112.3 ppb, with 

a maximum value of 365 ppb. Spirometry was performed by the children before and after a 

one-hour intermittent exercise session outdoors. Children were repeatedly tested up to 8 times. 

A 0.48% (95% CI: 0.24, 0.72) decline in FEV1 was experienced by the children after exercising 

outdoors. However, stratified analyses indicated that significant changes were observed only 

with higher quintiles of O3 exposure. At the highest quintile of exposure (183-365 ppb), 

a !2.85% (95% CI: !1.30, ! 4.40) change in FEV1 was observed postexposure. Therefore, 

children exercising outdoors when ambient O3 levels were high experienced declines in 

pulmonary function despite the repeated daily exposure to moderate and high levels of O3 in 

Mexico City. 

In the southern California Children’s Health Study, a total of 3,535 initially nonasthmatic 

children (ages 9 to 16 years at enrollment) were followed for up to 5 years to identify new-onset 

asthma cases (McConnell et al., 2002). Communities were stratified by pollution levels, with six 

high-O3 communities (mean 1-h max O3 of 75.4 ppb [SD 6.8] over four years) and six low-O3 

communities (mean 50.1 ppb [SD 11.0]). Asthma risk was not found to be higher for residents 

of the six high-O3 communities versus residents of the six low-O3 communities. However, 

within the high-O3 communities, asthma risk was 3.3 (95% CI: 1.9, 5.8) times greater for 

children who played three or more sports as compared with children who played no sports. This 

association was absent in the low-O3 communities (relative risk of 0.8 [95% CI: 0.4, 1.6]). 

Thus, similar to the results observed in the Mexico City study by Castillejos et al. (1995), greater 

effects were seen for susceptible individuals who exercised more outdoors at high O3 levels. 

The studies discussed above indicate that prolonged exposure periods, combined with 

elevated levels of exertion or exercise, may magnify the effect of O3 on lung function. Results 

from these studies are consistent with the earlier summer camp studies (Avol et al., 1990; 

Higgins et al., 1990; Raizenne et al., 1987, 1989; Spektor et al., 1988a, 1991) which also 

indicated large O3-related changes in lung function parameters in children who spent long hours 

outdoors. The large observed effects further suggest that other components in the ambient air 

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pollution mixture may potentiate the effects of O3. In conclusion, outdoor workers who spend, 

on average, 8 to 10 hours outdoors daily during the time of day when high peak O3 

concentrations are expected appear to be particularly vulnerable to O3 health effects and may 

need protection from O3 exposures. 

7.6.8 Summary of Key Findings and Conclusions Derived from Ozone 

Epidemiologic Studies 

In the previous 1996 O3 AQCD, there was considerable evidence of O3-related respiratory 

health effects from individual-level camp and exercise studies, as well as some consistent 

evidence from time-series studies of emergency room visits and hospitalizations. Since the 1996 

document, more field studies have been conducted, with some emphasis on additional outcome 

markers such as respiratory symptoms and asthma medication use. Another significant addition 

to the current O3 AQCD is the substantial number of short-term O3 mortality studies. The recent 

publication of an analysis examining the relationship between O3 and mortality in 95 U.S. 

communities (Bell et al., 2004) and three meta-analysis on O3-mortality associations (Bell et al., 

2005; Ito et al., 2005; Levy et al., 2005) also contribute significantly to the evidence base. 

Considering the wide variability in possible study designs and statistical model specification 

choices, the reported O3 risk estimates for the various health outcomes are in reasonably good 

agreement. In the case of O3-mortality time-series studies, combinations of choices in model 

specifications (e.g., the number of weather terms and degrees of freedom for smoothing of 

mortality-temporal trends) alone may explain the extent of the difference in O3 risk estimates 

across studies. As use of time-series studies to investigate air pollution effects has become more 

common, there has been a great effort to evaluate the issues surrounding these studies. 

The epidemiologic studies discussed in this chapter provide important information on the 

associations between health effects and exposures of human populations to ambient O3. 

A variety of oxidants in both the gaseous and particulate phases have not been examined in 

the available epidemiologic literature. The associations observed between ambient O3 

concentrations and health outcomes may represent O3 effects, per se, or O3 may be serving as a 

surrogate measure for the overall ambient photochemical oxidant mix. 

In this section, conclusions regarding O3 health effects from the epidemiologic evidence 

and the issues that may affect the interpretation of the effect estimates are briefly summarized. 

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A more integrative synthesis of all relevant information will be presented in Chapter 8 of 

this AQCD. 

(1) Field/panel studies of acute O3 effects. Results from recent field/panel studies 

continue to confirm that short-term O3 exposure is associated with acute decrements 

in lung function and increased respiratory symptoms, particularly in children and 

asthmatics. There is also suggestive evidence that O3 is related to increased asthma 

medication use. Strong O3 effects on lung function also were observed in outdoor 

workers. Taken together with the evidence from controlled human exposure studies, 

O3 is likely causally related to the various respiratory health outcomes. The current 

evidence is much more limited, though suggestive, of a potential effect of O3 on heart 

rate variability, ventricular arrhythmias, and the incidence of myocardial infarctions. 

(2) Acute O3 effects on emergency department visits and hospitalizations. Large multicity 

studies, as well as many studies from individual cities have reported an effect of O3 on 

respiratory hospital admissions. Studies using year-round data noted some 

inconsistencies; however, studies with data restricted to the summer or warm season, 

in general, indicated positive and robust associations between ambient O3 

concentrations and respiratory hospital admissions. Effects of O3 on asthma 

emergency department visits also were observed during the warm season. 

(3) Acute O3 effects on mortality. The majority of the studies suggest an elevated risk of 

all-cause mortality associated with acute exposure to O3, especially in the summer or 

warm season when O3 levels are typically high. Slightly greater O3 effects were 

observed for cardiovascular mortality. Results from recent large U.S. multicity timeseries studies provide the strongest evidence to-date for O3 effects on acute mortality. 

Recent meta-analyses also indicate positive risk estimates that are unlikely to be 

confounded by PM; however, future work is needed to better understand the influence 

of model specifications on the risk coefficient. 

(4) Chronic O3 exposure effects on morbidity and mortality. Fewer studies have 

investigated the effect of chronic O3 exposure on morbidity and mortality. The 

strongest evidence is for seasonal effects of extended O3 exposures on lung function in 

children, i.e., reduced lung function growth being associated with higher ambient O3 

levels. Longer-term studies investigating the association of chronic O3 exposure on 

yearly lung function, asthma incidence, and respiratory symptoms are inconclusive. 

Chronic O3-mortality studies observed inconsistent results across exposure periods and 

cause-specific mortality outcomes. 

(5) Exposure assessment. Exposure misclassification may result from the use of 

stationary ambient monitors to determine exposure in population studies. Although 

central ambient monitors do not explain the variance of individual personal exposures, 

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significant correlations are found between aggregate personal O3 measurements and O3 

concentrations from ambient monitors. A simulation study indicated that the use of 

ambient monitor data will tend to underestimate the O3 effect. A better understanding 

of the factors that affect the relationship between ambient concentrations and personal 

exposures will improve interpretation of the O3 effect estimates. 

(6) Ozone exposure indices. The three most commonly used daily O3 exposure indices, 

1-h max O3, 8-max O3, and 24-h avg O3, were found to be highly correlated in studies 

conducted in various regions. In addition, the effect estimates and significance of 

associations across all health outcomes were comparable when using the standardized 

distributional increment of 40 ppb, 30 ppb, and 20 ppb for mean 1-h max O3, mean 8-h 

max O3, and mean 24-h avg O3, respectively. 

(7) Lag structures for O3 exposure and effect. The lag time between O3 exposure and 

effect may differ depending on various factors such as the specific health outcome of 

interest, the mechanism of effect, and preexisting health conditions. The majority of 

the studies found an immediate O3 effect, with the strongest associations observed 

between health outcomes and O3 exposure on the same day and/or previous day. Some 

studies found large cumulative effects of O3 over longer lag periods, indicating that 

multiday lags also may be relevant for some health outcomes, including mortality and 

asthma symptoms in children. 

(8) Sensitivity to model specifications for temporal trends. Ozone effect estimates 

that were reported in studies whose main focus was PM often were calculated using 

the same model specifications as PM to adjust for temporal trends. While the 

sensitivity of the O3 risk estimates to alternative model specifications has not been 

throughly investigated, limited evidence indicates that O3 effects may be robust to 

various model specifications for temporal trend adjustment. 

(9) Sensitivity to model specifications for meteorological effects. Ozone risk estimates 

were generally more sensitive to alternative weather models than to varying degrees of 

freedom for temporal trend adjustment. In studies in which alternative weather models 

(e.g., quintile indicator model and four-smoother model) were considered, up to a 

factor of two difference in the O3 risk estimates was observed. Further research is 

needed to reduce the uncertainties related to confounding by weather influences. 

(10) Influence of seasonal factors. An evaluation of the confounding effects of 

meteorologic factors and copollutants on O3 risk estimates is complicated by their 

changing relationships with O3 across seasons. In addition, seasonal or seasonallymodified factors (e.g., air conditioning use, time spent outdoors) complicate 

interpretation of all-year effect estimates as they affect the relationship between 

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ambient concentrations and personal exposures. Given the potentially significant 

influence of season, season-specific analyses are more informative in assessing O3 

health risks. 

(11) Confounding by copollutants. Multipollutant regression models often are used to 

adjust for confounding by copollutants. Although there is some concern regarding the 

use of multipollutant models given the varying concurvity across pollutants, results 

generally suggest that the inclusion of copollutants into the models do not substantially 

affect O3 risk estimates. These findings indicate that effects of O3 on various health 

outcomes are robust and independent of the effects of other copollutants. 

(12) Concentration-response function. In the limited mortality and morbidity studies that 

have specifically examined the O3 concentration-response relationship, the evidence is 

inconclusive regarding the presence of an effect threshold. Factors such as exposure 

measurement error may reduce the ability to detect a threshold in population studies. 

The limited evidence suggests that if a population threshold exists in O3 health effects, 

it is likely near the lower limit of ambient O3 concentrations in the United States. 

(13) Heterogeneity of O3 health effects. Consistent O3 effect estimates generally were 

observed for mortality, hospitalizations, and other respiratory health outcomes in 

multicity studies. Some of the observed geographic heterogeneity in effects may be 

attributable to the differences in relative personal exposure to O3, which is affected by 

factors such as air conditioning prevalence and activity patterns, as well as varying 

concentrations and compositions of copollutants present by region. 

(14) Ozone health effects in asthmatics. The effects of O3 on asthmatics have been 

examined widely in both time-series studies and panel studies. Associations of 

O3 with various respiratory health outcomes, including lung function declines, 

increased respiratory symptoms, and emergency department visits, were observed. 

These findings, along with the pathophysiologic understanding of asthma as a chronic 

inflammatory disease, indicate that asthmatics are likely a susceptible population that 

requires protection from O3 exposures. 

(15) Age-related differences in O3 health effects. Supporting evidence exists for 

heterogeneity in the effects of O3 by age. The elderly population (>65 years of age) 

appears to be at greater risk of O3-related mortality and hospitalizations compared to 

all age populations. In addition, potentially adverse respiratory health outcomes are 

associated with O3 exposure in children (<18 years of age). Lung function responses 

to O3 exposure are diminished in the elderly population, as observed in an 

epidemiologic study and numerous human clinical studies. 

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8. INTEGRATIVE SYNTHESIS: OZONE EXPOSURE

AND HEALTH EFFECTS

8.1 INTRODUCTION 

This integrative synthesis is structured to provide a coherent framework for the assessment 

of health risks associated with human exposures to ambient surface-level (tropospheric) ozone 

(O3) in the United States. The main goal of the chapter is to integrate newly available scientific 

information with key findings and conclusions from the 1996 O3 AQCD (U.S. Environmental 

Protection Agency, 1996a), so as to address issues central to the EPA’s assessment of evidence 

needed to support the current review of the primary O3 NAAQS. The integrated assessment of 

key findings and conclusions provided here and elsewhere in this document with regard to O3 

exposure and health effects will be drawn upon and their policy implications considered in an 

Ozone Staff Paper prepared by EPA’s Office of Air Quality Planning and Standards (OAQPS). 

The analyses provided in that Staff Paper aim to “bridge the gap” between scientific assessments 

in this criteria document and judgments required of the EPA administrator in evaluating whether 

to retain or, possibly, to revise the current primary O3 NAAQS. Other types of scientific 

information concerning ambient O3 welfare effects (i.e., tropospheric O3 effects on vegetation 

and ecosystems, relationships to surface-level solar UV flux/climate changes, and effects on 

man-made materials) are assessed in ensuing Chapters 9, 10, and 11. That information will also 

be considered in the OAQPS staff paper in posing options regarding the secondary O3 NAAQS. 

As discussed in Chapter 2 of this document, O3 found in the earth’s troposphere generally 

originates from photochemical reactions that involve the interaction of sunlight with precursor 

pollutants, especially nitrogen oxides (NOx), carbon monoxide (CO), and volatile organic 

compounds (VOCs) such as hydrocarbons emitted by surface-level mobile and stationary 

sources and natural sources. Other photochemical oxidants, such as peroxyacetyl nitrate (PAN) 

and hydrogen peroxide (H2O2), are also generated along with O3 by such atmospheric 

interactions. In addition to the tropospheric O3 generated by these interactions, some O3 is found 

near the earth’s surface as the result of its downward transport from the stratosphere. However, 

in contrast to stratospheric O3, which plays an important role in maintaining the habitability of 

the planet by shielding the Earth’s surface from harmful solar ultraviolet (UV) radiation, 

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tropospheric O3 at the surface can also exert adverse effects on humans, nonhuman animal 

species, and vegetation. As was the case for previous O3-related NAAQS criteria revisions, the 

present criteria document focuses mainly on the assessment of health and welfare effects 

resulting from exposures to surface-level concentrations of tropospheric O3, whereas less 

attention is accorded to the distinctly much more limited available information on other 

photochemical oxidants, e.g., PAN or H2O2. 

Based on the criteria review completed in 1978, the original primary and secondary 

NAAQS set in 1971 for total photochemical oxidants were revised in 1979 to focus on O3 as the 

indicator for new primary and secondary standards that were attained when the expected number 

of days per calender year with daily maximum 1-h average O3 concentrations (1-h max O3) 

>0.12 ppm did not exceed one. The NAAQS for ambient O3 were revised in 1997 by replacing 

the 1-h standards with an 8-h primary standard that is met when the 3-year average of the annual 

fourth highest daily maximum 8-h average concentration (8-h max O3) is #0.08 ppm. The new 

1997 primary standard was based on various scientific supportive data from experimental human 

exposure studies, animal toxicologic studies, and epidemiological studies, as assessed in the 

1996 O3 AQCD and in the 1996 O3 Staff Paper (U.S. Environmental Protection Agency, 1996b). 

8.1.1 Chapter Organization 

In addition to providing the above brief background information regarding prior O3 

NAAQS reviews, this first section (8.1 Introduction) of the integrative synthesis chapter aims to 

orient the reader to the organization and content of the chapter. The next section (Section 8.2) 

focuses on air quality trends and current ambient O3 levels to help provide context for the 

ensuing discussions of O3 exposures and associated health effects. Subsequent sections 

then integrate newly available key scientific information assessed in Chapters 4 through 7 of 

this document, including integration of information on O3 dosimetry, toxicological information 

derived from controlled human exposure and laboratory animal studies, and epidemiologic 

evidence. 

These sections collectively address the following topics: (1) ambient O3 exposures, 

personal exposures, and dosimetric considerations; (2) experimental studies on toxicological 

responses to acute O3 exposures in humans (clinical studies) and both acute and chronic effects 

in animals; (3) epidemiological evidence for associations between ambient O3 exposure of 

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human populations and various health effects, as well as the strength and robustness of these 

associations; (4) integration of the experimental and epidemiological evidence; (5) biological 

mechanisms and other evidence useful in judging the plausibility of adverse health effects being 

associated with human exposures to ambient O3 levels encountered in the United States; and 

(6) delineation of susceptible and vulnerable populations likely at increased risk for O3-related 

health effects and numbers of people potentially falling in such categories in the United States. 

8.2 AMBIENT OZONE AIR QUALITY IN THE UNITED STATES 

8.2.1 Current Ozone Concentrations and Spatial Patterns 

The EPA has established “O3 seasons,” during which ambient O3 concentrations must be 

monitored in the United States and its territories. These seasons vary in length depending on 

location. The O3 season extends all year in the Southwest. In most other areas of the country, O3 

is monitored typically from April to October. However, O3 is monitored throughout the year in 

many urban areas, because O3 is present year round not only in polluted areas but in clean areas 

as well. The median of the daily 8-h max O3 in the United States, averaged over May to 

September from 2000 to 2004 for all U.S. counties, was 0.049 ppm. In 95% of all counties, the 

median of the daily 8-h max O3 was less than 0.057 ppm. However, it should be noted that most 

monitors are located in the East and O3 data are sparse throughout large areas of the West. 

Median values of daily 1-h max O3 were typically much higher in large urban areas or in areas 

downwind of them. For example, in Houston, TX they approached 0.20 ppm during the same 

2000-2004 period. Daily 1-h max O3 concentrations were lower in the rest of the country, but 

were still above 0.12 ppm in many locations. Eight-hour daily maximum concentrations were 

not as high, but tend to be highly correlated with 1-h daily maximums. 

Within individual Metropolitan Statistical Areas (MSAs), O3 concentrations tend to be well 

correlated across monitoring sites, although spatial variations in concentrations can be 

substantial. In many city centers, O3 concentrations tend to be lower than in either upwind or 

downwind areas, largely due to reaction of O3 with NO emitted by motor vehicles. For example, 

much lower O3 concentrations overall are found in downtown Los Angeles (e.g., in Lynwood) 

than at sites located further downwind (e.g., in San Bernadino). The much higher downwind 

levels are formed from photochemical reactions involving the urban emissions, including 

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products formed as the result of reactions titrating O3 in the urban core. Thus, O3 concentrations 

tend to be higher downwind of urban centers, and they decrease again in areas that are more 

remote from precursor sources. Likewise, surface-level O3 can be depleted in rural areas close to 

NO emission sources, such as highways and powerplants. 

8.2.2 Diurnal and Seasonal Variations 

Ozone concentrations typically tend to peak in early to mid-afternoon in areas where there 

is strong photochemical activity and to peak later in the afternoon or during early evening in 

areas where transport is more important in determining the O3 abundance. Summertime maxima 

in O3 concentrations occur in those U.S. areas where substantial photochemical activity acts 

on O3 precursors emitted as the result of human activities. Monthly maxima can occur anytime 

from June through August. However, springtime maxima are observed in some National Parks, 

mainly in the western United States, and at a number of other relatively unpolluted monitoring 

sites throughout the Northern Hemisphere. For example, the highest O3 concentrations at 

Yellowstone National Park tend to occur during April and May. Typically, monthly minima 

tend to occur from November through February at polluted sites and during the fall at relatively 

remote sites. 

8.2.3 Long-Term Trends 

National attention started to be focused in the 1940s on O3 and associated photochemical 

smog in the Los Angeles area. Prior to the adoption of stringent emissions controls to reduce O3 

precursors, peak O3 levels were consistently higher in the Los Angeles area than are currently 

observed. For example, in 1958, peak O3 concentrations measured in Los Angeles were about 

0.6 ppm but have declined since then, although not at a steady rate. Peak O3 levels of 0.2 to 

0.5 ppm were still found at some locations in the Los Angeles basin during the 1970s. 

For example, on two days (October 13 and 14) during a 1978 episode, Tuazon et al. (1981) 

observed peak 1-h averaged O3 values of nearly 0.4 ppm and nearly 0.5 ppm. Currently, peak 

1-h and 8-h average O3 concentrations are about 0.17 and 0.15 ppm in the Los Angeles basin 

(cf., Figures 3-10 and 3-11). High O3 levels were also earlier found throughout the rest of the 

United States as well, but peak O3 levels have also gradually declined across the country during 

the 1980s and 1990s. However, during one particularly hot summer (of 1988) in the East, peak 

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1-h O3 concentrations of about 0.2 ppm were observed in many eastern U.S. cities (U.S. 

Environmental Protection Agency, 1990). 

Nationwide, 2nd highest 1-h ozone concentrations in the United States have decreased 

dramatically during the past several decades, i.e., by ~29 percent from 1980 to 2003 and 

16 percent from 1990 to 2003. Also, 4th highest 8-h O3 concentrations decreased by ~21 percent 

since 1980 and 9 percent since 1990 (U.S. Environmental Protection Agency, 2003). Trends in 

metrics for evaluating compliance with the O3 NAAQS (i.e., changes in the 4th highest O3 

concentration) can be found in EPA’s “National Air Quality and Emissions Trends Reports.” 

These reports indicate that the 4th highest O3 concentrations are still decreasing nationwide, but 

the rate of decrease has slowed since 1990. However, such trends have not been uniform across 

the United States. In general, reductions in the O3 metrics given above have been largest in New 

England and in states along the West Coast and smallest in midwestern states. Downward trends 

in California O3 concentrations have been driven mainly by notable decreases in Southern 

California, with smaller reductions occurring in other areas. Trends in peak O3 metrics do not 

necessarily reflect changes in O3 concentrations across the middle of the distribution of ambient 

O3 values. Of note, O3 concentrations towards the center of its nationwide distribution have not 

shown much change (cf., Figures 3-17, 3-18, and Table AX3-9), and there are some indications 

that O3 concentrations at the lower end of the distribution may even be increasing. 

8.2.4 Interrelationships Between Ozone and Other Ambient Pollutants 

Data on ambient concentrations of other oxidants (e.g., H2O2, PAN) and oxidation products 

(e.g., HNO3, H2SO4) in the atmosphere are not nearly as abundant as they are for O3. Because 

data for such species are usually obtained only as part of specialized field studies, it is difficult to 

relate observed ambient O3 concentrations to ambient levels of other oxidant species or oxidation 

products. In general, such secondary species are expected to be at least moderately positively 

correlated with O3, whereas, primary species are expected to be more highly correlated with each 

other than with secondary species (provided that the primary species originate from common 

sources in given areas). Measurements of gas phase oxidants obtained as part of the Southern 

Oxidants Study (SOS) found combined hydroperoxide (H2O2, CH3OOH, and HOCH2OOH) 

concentrations typically in the range of several ppb. Other SOS measurements likewise 

indicated combined concentrations of PAN, PPN, and MPAN in the range of several ppb. 

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Oxidants are also present in airborne cloud droplets, rain drops, and particulate matter (PM). 

A few measurements of reactive oxygen species (ROS), expressed as equivalent H2O2, 

in ambient fine PM have indicated levels of less than 1% of those for ambient O3 on a molar 

basis. However, it should be noted that these measurements are potentially subject to both 

positive and negative artifacts. 

PM

Relationships between ambient air O3 and PM levels can be quite complex because PM 

is not a single distinct chemical species, but rather a mix of primary and secondary species. 

As an example of this complexity, O3 concentrations were found to be positively correlated with 

2.5 during the summer, but negatively correlated during the winter at Ft. Meade, MD (cf., 

Figure 3-21). Also, Ito et al. (2005) examined relationships between ambient O3 and PM10 on a 

seasonal basis in several U.S. urban areas (cf., Figure 7-24). Seasonal O3-PM relationships 

similar to those observed at Ft. Meade were found, reflecting the dominant contribution of PM2.5 

to PM10 in the urban areas studied. Possibly contributing to higher correlations seen between O3 

and fine PM in the summer is the fact that O3 can contribute to formation of submicron particles 

via interactions with various other atmospheric constituents present, such as terpenes and other 

biogenically derived hydrocarbons from trees, other vegetation, and wood products. Formation 

of ultrafine particles by this mechanism mostly occurs during summer days when temperatures 

and O3 concentrations are sufficiently elevated to facilitate O3 reactions with increased amounts 

of biogenic hydrocarbons emitted from vegetation. Bursts of ultrafine particle formation have 

been observed repeatedly in both urban and rural air. Woo et al. (2001), for example, reported 

rapid formation of ultrafine particles in the ambient air of Atlanta typically around noon in both 

summer and winter. The mechanisms underlying such ultrafine particle formation events may 

also involve other atmospheric reactions related to O3 formation, such as nucleation of H2SO4 

(produced by oxidation of SO2) and, probably, NH3. 

8.2.5 Policy Relevant Background (PRB) Ozone Concentrations 

Policy relevant background (PRB) O3 concentrations, i.e., background O3 concentrations 

used for NAAQS-setting purposes, are those that would occur in the United States in the absence 

of anthropogenic emissions in continental North America (defined here as the United States, 

Canada, and Mexico). Such PRB O3 concentrations include contributions from natural sources 

everywhere in the world and from anthropogenic sources outside these three countries. For the 

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purpose of informing O3 NAAQS decisions, EPA focuses on assessing risks to human health and 

environmental effects from O3 levels in excess of PRB concentrations. Issues concerning the 

methodology for estimating PRB O3 concentrations are discussed in detail in Annex AX3, 

Section AX3.9. 

Contributions to PRB O3 include photochemical reactions involving natural emissions of 

VOCs, NOx, and CO, as well as the long-range transport of O3 and its precursors from outside 

North America and the stratospheric-tropospheric exchange (STE) of O3 (see Annex AX2, 

Section AX2.3 for details regarding O3-related STE processes). Natural sources of O3 precursors 

include biogenic emissions, wildfires, and lightning. Biogenic emissions from agricultural 

activities are not considered in the formation of PRB O3. However, estimates of PRB O3 

concentrations cannot be derived solely from measurements of O3 at relatively unpolluted sites 

because of long-range transport from anthropogenic source regions within North America. It is 

impossible to determine sources of O3 at a particular location without ancillary data that could be 

used as tracers of sources or to calculate photochemical production and loss rates for O3. Thus, 

estimates of PRB O3 concentrations are currently based on predictions generated by the global 

scale, three dimensional, chemical transport model GEOS-CHEM (Fiore et al., 2003). 

Policy relevant background O3 concentrations vary as a function of season, altitude, and 

total surface O3 concentration, with PRB O3 concentrations at the surface generally falling in the 

range of 0.015 to 0.035 ppm from 1300 to 1700 local time and tending to decline under weather 

conditions conducive to O3 episodes. The PRB concentrations are highest during spring and 

decline into summer; and higher values also tend to occur at higher elevations during the spring 

due to contributions from hemispheric pollution and stratospheric intrusions. The contribution to 

surface O3 by stratospheric intrusions is typically well below 0.020 ppm. Stohl (2001) and 

Sprenger et al. (2003) found that the maximum probability of stratospheric intrusions reaching 

the 800 hPa level (~1800 m) was less than 1% and that higher probabilities (1 to 2%, and 10%) 

applied for stratospheric intrusions penetrating to the 600 hPa level (~4100 m) and 500 hPa level 

(~5400 m), respectively. Thus, stratospheric O3 intrusions only rarely contribute to elevated 

surface-level O3 concentrations at low altitude sites but have a higher (albeit still low) 

probability of elevating them at high-altitude sites. 

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8.3 FACTORS AFFECTING HUMAN EXPOSURE TO AMBIENT OZONE 

Human exposure to O3 and related photochemical oxidants varies over time due to changes 

in their ambient concentrations and because people move between locations having notably 

different O3 concentrations. Also, the amount of O3 delivered to the lung is not only influenced 

by the ambient concentration but also by the individual’s breathing route and rate. Thus, activity 

level is an important consideration in determining potential O3 exposure and dose received. 

The use of data from ambient air monitoring stations is still the most common surrogate for 

assigning exposure estimates in epidemiologic studies. Since the primary source of O3 exposure 

is the ambient air, O3 concentration data from outdoor community monitoring sites should 

provide a relative assignment of exposure with time, if O3 concentrations are relatively uniform 

across the region; time-activity patterns are roughly the same across the study population; and 

housing characteristics (such as ventilation rates and O3 sinks contributing to indoor O3 decay 

rates) are relatively constant for the study area. However, because these types of factors often do 

vary across populations and locations, some error tends to be associated with estimates of the 

magnitude of O3 exposure of large populations. Nevertheless, although substantial variability 

may exist among personal measurements, human exposure studies have observed that daily 

average personal O3 exposures for the general population tend to be reasonably well correlated 

with monitored ambient O3 concentrations. Therefore, ambient O3 monitoring data appear to 

provide the most useful index of human O3 exposure currently available to help characterize 

health outcomes associated with O3 exposures of large population groups. 

8.3.1 Personal Exposure 

Personal O3 concentrations have been measured for children, outdoor workers, and 

individuals with pulmonary diseases (populations potentially vulnerable to increased ambient O3 

exposure and/or susceptible to O3 or other respiratory irritants). Outdoor workers can be 

expected to have somewhat higher O3 exposures than other individuals, because they typically 

spend more time outdoors and often engage in prolonged moderate and heavy exertion activities. 

Children also tend to be more active outside and, therefore, often manifest a higher breathing 

rate than most adults. Given the higher percentages of time spent outdoors than most other 

population groups, personal exposures of both outdoor workers and children tend to be more 

highly correlated with ambient O3 concentrations measured at community monitoring sites. 

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8.3.2 Indoor Concentrations 

Apart from only a few specific indoor sources such as photo-copying machines, O3 indoors 

is derived from the infiltration of ambient air from outdoors. Generally, O3 enters indoor 

environments through infiltration from the outdoors and through building components, such as 

windows, doors, and ventilation systems. Ozone concentrations in indoor environments depend 

primarily on the outdoor O3 concentration, outdoor/indoor infiltration and the air exchange rate 

(AER). Hence, indoor O3 concentrations tend to reflect outdoor concentrations and are higher 

when outdoor O3 is higher. 

Once indoors, O3 reacts on various surfaces and with airborne components of either indoor 

or outdoor origin. Because O3 reacts indoors with surfaces and other contaminants, O3 

concentrations are typically lower indoors than outdoors. Gas phase reactions occurring indoors 

also produce other oxidants analogous to the production of photochemical smog. The extent and 

rate of production of these other species indoors is a function of indoor O3 concentrations and the 

presence of other necessary precursors (i.e., VOCs), along with an optimal AER. 

Several studies of O3 measured in residences, schools, office buildings and museums found 

that typical O3 concentrations varied across all such locations. Indoor O3 levels generally varied 

in relationship to the AER in the indoor environment (increasing with higher AER) and were 

generally notably lower than outdoor ambient O3 levels. For example, one study of O3 levels 

indoors and outside of a school in New England found average O3 concentrations of 40 ppb 

(0.040 ppm) outdoors and 20 ppb (0.020 ppm) indoors. With regard to O3 levels in mobile 

source microenvironments, as is the case for other enclosed environments, O3 concentrations 

depend on the extent of mixing of outdoor air into the vehicle cabin. Thus, if windows are kept 

open, O3 concentrations inside the vehicle may approach outdoor values. But, if windows are 

kept closed and there is air conditioning, then interior values can be much lower than those 

outside, especially if recirculated air is used. For example, in one N.C. study of police cars with 

air conditioning and recirculated air, vehicle cabin O3 levels (11.7 ppb average) were less than 

half those outside (28.3 ppb average at area outdoor monitoring sites). 

Although O3 concentrations may be reduced to lower levels once ambient O3 enters 

indoor environments, the indoor O3 may interact with other airborne substances of indoor or 

outdoor origin present indoors. For example, Wainman et al. (2000) found that O3 reacts with 

d-limonene, a common component of air fresheners to produce submicron particles found mainly 

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to range in size from 0.1 to 0.3 μm. Wainman et al. noted that terpenes such as limonene are 

emitted by wood products; that they are used as solvents, as odorants in cleaning products, and 

as air fresheners; and, even though they are produced by vegetation outdoors, their indoor 

concentrations are often higher than outdoors because of their widespread uses. In addition to 

particle formation, Weschler (2004) noted that gas phase products (i.e., aldehydes and 

hydroperoxides) produced by O3 reactions with terpenes and other unsaturated carbon 

compounds may also be of concern. During formation of such products, OH radicals are also 

generated that can react with compounds that do not react with O3. To the extent that building 

ventilation rates remain constant between days characterized by high and low ambient O3, the 

concentrations of these other secondary pollutants formed indoors will tend to be correlated with 

ambient O3. Thus, ambient O3 concentrations measured outdoors at community monitoring sites 

and/or personal O3 exposure monitor measurements may serve not only as indices of direct 

human exposure to O3 per se, but also as surrogate indices of exposures to broader O3-containing 

ambient mixtures of photochemical oxidants and/or other pollutants. 

8.4 SYNTHESIS OF AVAILABLE INFORMATION ON OZONERELATED HEALTH EFFECTS 

The integrated synthesis of the latest available information on O3-related health effects 

poses an interesting challenge in view of the emergence of highly important new information 

since the 1996 O3 AQCD. Such information includes new findings from: 

C Dosimetry studies that further characterize factors potentially affecting regional 

distribution of O3 in the respiratory tract of humans and laboratory animals and provide 

improved bases for animal-to-human extrapolations of experimentally-observed 

O3-induced health effects. 

C Experimental toxicological studies using controlled human exposures and laboratory 

animals aimed at delineating exposure-response relationships and understanding potential 

biochemical mechanisms underlying toxic effects, pathology, and susceptibility. 

C Epidemiological studies, reflecting progress in addressing many research needs identified 

during the last review, as well as raising new issues and reevaluating previously addressed 

issues that remain important in interpreting the body of epidemiological evidence and 

characterization of its strengths and limitations. 

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Previous criteria assessments, including the 1996 O3 AQCD, found that experimental 

studies of controlled human and laboratory animal exposures to O3 provided the most clear cut 

and compelling evidence with regard to characterizing O3-related health effects. Based on 

extensive dosimetric and experimental data, as well as growing epidemiologic evidence 

available at the time, the 1996 O3 AQCD arrived at a set of findings and conclusions regarding 

potential health effects of ambient O3 exposure. In general, the existing evidence was such to 

warrant a high degree of confidence in those conclusions derived from experimental (controlled 

exposure) studies of acute O3 exposure effects. Considerable confidence could also be placed in 

the emerging field/panel studies providing observational study results substantiating and 

extending the controlled exposure study findings. The controlled exposure and field/panel 

studies clearly demonstrated respiratory effects (e.g., reduced pulmonary function) among 

healthy and asthmatic children and adults acutely exposed for 1 to 8 h to 0.08 ppm O3 while 

physically active. Other epidemiologic studies provided highly suggestive, although less 

conclusive, indications of increased morbidity (e.g., as indexed by emergency department [ED] 

visits, hospital admissions, etc.) and, possibly, mortality being associated with acute exposure of 

human populations to ambient O3. Also, data from laboratory animal and controlled human 

exposure studies supported the hypothesis that coexposure to O3 and other pollutants at loweffect levels can result in enhanced effects. However, the issue of exposure to copollutants was 

poorly understood, especially with regard to potential chronic effects. 

Since the 1996 O3 AQCD evaluations, further controlled human exposure studies have 

extended earlier findings of respiratory effects of acute exposures in exercising adults to O3 

concentrations ranging below 0.08 ppm for some sensitive subjects. Also, a more extensive 

database of air pollution epidemiologic studies has become available; and a subset of these new 

observational studies have reported a variety of O3-related health effects associations, with 

newly reported evidence of ambient O3 mortality relationships being of special interest. Based 

on physiological, biochemical, and molecular changes observed in controlled human exposure 

studies and animal toxicological studies, new evidence is now available by which to more fully 

evaluate the biological plausibility and extent of coherence for various health outcomes (such as 

respiratory and cardiovascular effects and mortality) reported in epidemiologic studies. 

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8.4.1 Integration of Experimental and Epidemiologic Evidence 

8.4.1.1 Cross-Cutting Issues Relevant to Assessment/Interpretation of Ozone 

Health Effects 

Three chapters in the current document provide detailed discussion of various experimental 

approaches utilized to evaluate O3-related health effects. Chapter 4 discusses dosimetry issues 

pertinent to both animal and human exposure scenarios. Chapter 5 discusses the experimental 

studies of biochemical (cellular and molecular changes), physiological, and pathological 

observations in laboratory animals (including nonhuman primates, dogs, and rodent species) 

and in vitro studies using cell culture systems, in certain cases, on humans cells recovered from 

bronchoalveolar lavage fluid (BALF) after exposure to O3. Chapter 6 evaluates studies on 

human volunteers exposed to O3 which have investigated a variety of physiological and 

biochemical endpoints. In interpreting the results from the experimental approaches, one must 

consider the following issues: (1) exposure/dose considerations; (2) interpretation of results for 

animals and humans from relatively high dose O3 exposures; (3) animal-to-human 

extrapolations, and (4) experimental results derived from single pollutant exposures in 

comparison to much more complex ambient exposures. 

Earlier animal toxicology studies were carried out using relatively high O3 exposure 

concentrations/doses that do not necessarily reflect “real-world” exposure scenarios. Those 

experiments were primarily aimed at understanding the pathophysiology associated with O3 

exposure in healthy animals, to help identify potential mechanisms(s) of action and to help 

validate health outcomes reported in epidemiologic studies. Since the 1996 O3 AQCD, some of 

the human and animal studies have used ambient and/or near ambient doses. The majority of 

controlled chamber exposure studies on human volunteers mainly limited exposures to O3 alone 

in comparison to sham (clean air) exposures, thus providing evidence concerning direct effects 

of O3 per se versus more closely mimicking real-world atmospheric exposures to multipollutant 

mixes. Some limited data from laboratory animal and controlled human exposure studies 

reviewed in the 1996 O3 AQCD pointed towards enhancement of some observed health effects 

by coexposures to O3 and other air pollutants. Since then a few newer human clinical and 

laboratory animal air pollution studies have utilized various coexposure regimens to simulate 

more closely ambient exposure to air pollution mixtures; and the results from these studies are 

highly useful in developing better models to interpret toxicologic effects associated with 

O3-containing ambient air pollutant mixes. 

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Interpretations of experimental studies of air pollution effects in animals, as in the case of 

environmental comparative toxicology studies, are affected by limitations associated with 

extrapolation models. The differences between humans and rodents with regard to O3 absorption 

and distribution profiles based on breathing pattern, exposure dose, and differences in lung 

structure and anatomy (see Chapter 4 and 5 for details) all have to be taken into consideration. 

Also, in spite of a high degree of homology and the existence of a high percentage of 

orthologous genes across human and rodents (particularly mice), extrapolation of molecular 

alterations at the gene level is complicated by species-specific differences in transcriptional 

regulation. Given these molecular differences, there are large uncertainties associated with 

quantitative extrapolations at this time between laboratory animals and humans of observed O3­

induced pathophysiological alterations under the control of widely varying biochemical, 

endocrine, and neuronal factors. 

Epidemiologic studies provide important additional observational information to our 

knowledge base regarding O3-induced health effects. Of particular importance, such studies help 

(1) to ascertain the extent to which the types of health effects found in experimental studies 

occur in response to ambient exposures of human populations to O3 and/or other copollutants; 

and (2) to delineate exposure-response relationships for more serious health effect outcomes 

(e.g., hospital admissions, mortality) that help to elucidate more fully the likely public health 

consequences of ambient O3 exposures. 

The ensuing sections both briefly summarize key dosimetry and health related findings 

derived from the 1996 O3 AQCD and integrate those findings with new information obtained 

since 1996 from (a) human and animal experimental studies and (b) epidemiologic studies. 

Physiological, pathological, cellular and biochemical alterations induced by experimental O3 

exposures are evaluated and compared to human health effects associated with ambient O3 

exposures. Also, the influence of O3-induced changes at cellular and molecular levels are 

integrated to help elucidate mechanistic bases for the observed physiological and pathological 

alterations. These research results are evaluated both to help assess the biological plausibility of 

health outcome associations observed in epidemiologic studies and to assess the coherence of the 

overall body of evidence relevant to O3-related health outcome conclusions, to aid in drawing 

conclusions about the likelihood of causal relationships existing between ambient exposure to O3 

and various types of reported O3-related health effects. 

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8.4.1.2 Dosimetry 

The state-of-the-art of O3 dosimetry, as described in 1996 O3 AQCD, indicated consistency 

across data and models derived from in vivo human and animal studies, thus increasing the level 

of confidence in the development of dosimetric extrapolation models. Earlier dosimetry models 

predicted that the tissue dose of inhaled O3 was greatest at the bronchoalveolar junction, the lung 

region experimentally shown to be most impacted by O3. Ozone bolus inhalation studies in 

humans had indicated that inspired O3 reaches the distal airways and alveoli of resting humans; 

and, with increased inspiratory flow rates due to exercise, O3 penetrates deeper and in greater 

quantity to the distal regions of the lung. These findings were corroborated by observations of 

18O3 (oxygen-18-labeled ozone) in the BALF of humans and rats (Hatch et al., 1994). Based on 

BALF parameters monitored in this study, an exposure of 0.4 ppm O3 in exercising humans 

appears to be approximately equivalent to an exposure of 2 ppm in resting rats. 

Some O3-induced acute responses compared well across species when controlled for dose, 

indicating that animals and humans (a) respond to O3 in a dose-dependent manner, i.e., they 

exhibit increasing breathing frequency with an accompanying decrease in tidal volume 

(tachypnea), and (b) show similar changes in alveolar permeability as measured by protein in the 

BALF. These parallel changes in humans and animals were sufficiently homologous to suggest 

a common mode of action. The majority of lung function decrements were seen to subside with 

repeated exposures in both humans and animals, with analogous attenuation of certain (but not 

all) parameters measured in the BALF. The mechanisms associated with attenuation are unclear 

but may involve endogenous antioxidants. The significance of non-attenuated markers in BALF 

has been interpreted to relate to potential chronicity of O3 effects. 

During the past decade, no further reports have been published on O3 uptake studies in 

animals, although several controlled human bolus and/or general O3 uptake studies have 

provided refined data. The bolus uptake studies suggest that prior exposure to O3 diminishes 

bolus uptake. In the 1996 O3 AQCD, the effect of mode of breathing (oral or nasal) on O3 uptake 

was thought to be minimal, with approximately equal uptake via the nose or mouth. Newer 

bolus dose studies have demonstrated that the uptake and regional respiratory tract distribution 

of O3 is sensitive to mode of breathing (nasal uptake greater than oral) and to air flow rate 

(uptake decreases with increasing flow). Studies (Rigas et al., 1997, Bush et al., 2001) that 

evaluated the role of co-exposure with other oxidants such as NO2 and SO2 recognized the 

8-14
importance in mediating O3 toxicity of other oxidative environment components in the lung, 

e.g., radicals generated by O3 interaction with the lipids in the airway and epithelial lining fluid 

(ELF). 

New uptake studies (Ultman et al., 2004) carried out in controlled human clinical studies 

have observed gender-specific differences in the uptake of O3, but these differences do not 

correlate well with spirometric responses. Rather, they appear to be related to breathing pattern 

and lung size, due to females having smaller lungs than males. Thus, a number of variables 

seem to affect O3 uptake, notably including route of breathing, breathing pattern, gender, 

copollutants, and certain pre-exposure conditions. These differences are important in order to 

interrelate experimentally-demonstrated pathophysiological effects and epidemiologicallyobserved associations between ambient O3 concentrations and health risks among human 

population groups. 

8.4.2 Experimental Evidence for Ozone-Related Health Effects 

Pulmonary Function 

Numerous controlled human exposure, animal, and epidemiological studies assessed in the 

1996 O3 AQCD demonstrated alterations in various pulmonary function measures. Ozone 

inhalation for several hours (1 to 3 h) while physically active was shown to elicit both acute 

pathophysiologic changes and subjective respiratory tract symptoms. The pulmonary responses 

observed in healthy human subjects exposed to ambient O3 concentrations included decreased 

inspiratory capacity; mild bronchoconstriction; rapid, shallow breathing pattern during exercise; 

and subjective symptoms of tracheobronchial airway irritation, including cough and pain during 

inspiration. Acute O3 exposures were also found to cause decreases in forced vital capacity 

(FVC), forced expiratory volume in 1 s (FEV1), peak expiratory flow (PEF), and increased 

airways resistance (SRaw). The severity of symptoms and the magnitude of pulmonary response 

were reported to vary as a function of exposure concentration, duration, and level of exercise 

(which help to determine inhaled O3 dose), as well as the individual sensitivity of exposed 

subjects and the extent of tolerance resulting from previous exposures. With regard to the latter, 

the 1996 O3 AQCD noted that during repeated short-term exposures, some of the O3-induced 

responses (inflammation and lung function decrements) are partially or completely attenuated 

with a differential attenuation profile. Over a 5-day exposure, both pulmonary and inflammatory 

8-15
markers were found to attenuate by the fifth day of exposure, but markers of cell damage 

(discussed below under Lung Injury) do not attenuate and continue to increase. 

Group mean data from numerous controlled human exposure (in healthy subjects 8 to 

45 years of age, after 1 to 3 h exposure) and field studies indicated that, in general, statistically 

significant pulmonary function decrements beyond the range of normal measurement variability 

(e.g., 3 to 5% for FEV1) occur (a) at >0.50 ppm O3 when at rest; (b) at >0.37 ppm O3 with light 

exercise (slow walking); (c) at >0.30 ppm O3 with moderate exercise (brisk walking); (d) at 

>0.18 ppm O3 with heavy exercise (easy jogging) and (e) at >0.16 ppm O3 with very heavy 

exercise (running). Smaller group mean changes (e.g., <5%) in FEV1 occurred at lower 

O3 concentrations with very heavy exercise in healthy adults at 0.15 to 0.16 ppm O3 and in 

healthy young adults at levels as low as 0.12 ppm. Also, pulmonary function decrements were 

seen in children and adolescents at O3 concentrations of 0.12 and 0.14 ppm with heavy exercise. 

Human studies reviewed in the 1996 O3 AQCD that used longer duration (6- to 8-h) acute 

exposures with exercise (which better mimic multihour exposures to ambient O3 that typify more 

prolonged elevated ambient O3 levels often observed in U.S. urban areas) provided some of the 

strongest and most quantifiable concentration-response data on pulmonary function effects of 

acute O3 exposure. Overall, the 1996 O3 AQCD found that for healthy subjects performing 

moderate exercise during longer duration (6 to 8 h) acute O3 exposures, group mean 5% 

decrements in FEV1 were seen at relatively low O3 levels, such as: 0.08 ppm O3 after 5.6 h, 

0.10 ppm O3 after 4.6 h, and 0.12 ppm O3 after 3 h. Some subjects in the earlier studies 

experienced FEV1 decrements in excess of 15% with short-term (1- to 3-h) or longer-duration 

(6.6-h) acute O3 exposures, suggesting potential wide interindividual variability in pulmonary 

function responses. 

The few newly available controlled human exposure studies (since those assessed in the 

1996 O3 AQCD) that used near-ambient O3 concentrations (#0.12 ppm) are summarized in 

Appendix Table 8A-1; and these, like most controlled exposure studies to date, continue to 

indicate that considerable interindividual differences exist in the magnitude of responses to O3. 

However, a given individual’s lung function and, to a lesser extent, respiratory symptom 

responses to O3 are reproducible over a period of time, indicating that some individuals are 

consistently more responsive than others to O3. Figure 8-1A illustrates well the variability in 

FEV1 responses in young healthy adults following a prolonged (6.6 h) exposure to O3, as 

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Figure 8-1A,B. Frequency distributions of FEV1 changes following 6.6-h exposures to a 

constant concentration of O3 or filtered air. During each hour of the 

exposures, subjects were engaged in moderate exercise for 50 minutes. 

With increasing O3 concentration, the distribution of responses becomes 

asymmetric, with a few individuals exhibiting large FEV1 decrements. 

Note that the percentage in each panel indicates the portion of subjects 

having a FEV1 decrement in excess of 10%. 

Source: Panel A, McDonnell (1996); Panel B, Adams (2002, 2006), pre- and post-FEV1 data for each subject 

provided by author. 

summarized in the 1996 O3 AQCD. Referring to this figure, the group mean FEV1 response 

following continuous (square-wave) exposure to 0.08 ppm O3 was relatively small (between a 

5 and 10% decrement). However, 18% of the exposed subjects had moderate FEV1 decrements 

of 10 to 20%, and 8% experienced large FEV1 decrements of greater than 20%. This serves to 

emphasize that, while group mean responses may be small and seem physiologically 

insignificant, some individuals can experience distinctly larger effects under similar O3 exposure 

conditions. Newer data from Adams (2002, 2006), as illustrated in Figure 8-1B, demonstrate 

notable interindividual variability for O3 exposure concentrations at and even below 0.08 ppm. 

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Following 6.6-h continuous exposure to 0.08 ppm O3 under intermittent, moderate exercise 

conditions, the group mean FEV1 decrement was 5%, but 17% of the subjects had greater than a 

10% decrease in FEV1. Following exposure to 0.06 ppm O3, the group mean FEV1 decrement 

was less than 2%. However, five subjects still had greater than a 5% decrease in FEV1 whereas 

only one experienced this magnitude of effect following exposure to filtered air. 

It should be noted that spirometry typically improves in healthy young adults with exercise 

exposures to filtered air (FA). The term “O3-induced” was used in Chapter 6 and Annex 6 to 

designate effects that have been corrected for these filtered air responses. For healthy adults, 

an O3-induced change in lung function is the difference between the decrement experienced 

with O3 exposure and the improvement observed with filtered air exposure. The FEV1 responses 

illustrated in Figure 8-1 were not corrected for the responses following filtered air exposures. 

For comparison to Figure 8-1B, O3-induced FEV1 responses from the Adams (2002, 2006) 

studies are illustrated in Figure 8-2. For the exposures to 0.04, 0.06, and 0.08 ppm O3, 

an O3-induced FEV1 decrement of $10% was experienced by 7, 7, and 23% of subjects, 

respectively. Effects of a 0.04 ppm O3 exposure were not apparent when simply comparing 

pre-post FEV1 (Figure 8-1B), whereas they are evident when considering O3-induced FEV1 data 

(Figure 8-2). The distinction between an O3-induced change and a post- versus preexposure 

change is particularly important in individuals with respiratory disease who may experience 

exercise-induced decrements in pulmonary function during both filtered air and O3 exposures. 

Other new studies (assessed in Chapter 6 and Annex 6 of this document) that evaluated 

responses in hundreds of subjects clearly indicate that O3-related FEV1 decrements and symptom 

responses decrease with age beyond young adulthood (18 to 20 years). For example, Hazucha 

et al. (2003) studied possible gender and age differences in O3 responsiveness and found that 

young females lose O3 sensitivity faster than young males, but the rate is about the same for both 

genders by middle age. 

A few controlled human exposure studies that explored a triangular exposure profile of 

average O3 concentrations between 0.08 to 0.12 ppm over 6.6 to 8 h (to more closely mimic the 

typical ambient O3 exposure pattern) observed greater overall FEV1 decrements with triangular 

exposures compared to constant or square-wave exposures (see Appendix Table 8A-1 and Annex 

Figure AX6-3). Furthermore, the peak FEV1 decrements observed during triangular exposures 

exceed those observed during square-wave exposures. At a lower average O3 concentration of 

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Figure 8-2. Frequency distributions of FEV1 changes following 6.6-h exposures to a 

constant concentration of O3 or filtered air. The FEV1 changes following O3 

exposures have been corrected for filtered air responses, i.e., they are O3­

induced FEV1 changes. These data should be compared with Figure 8-1B. 

Note that the percentage in each panel indicates the portion of subjects tested 

having FEV1 decrements in excess of 10%. 

Source: Adams (2002, 2006), pre- and post- FEV1 data for each subject provided by author. 

0.06 ppm, no temporal (i.e., hour by hour responses) differences were observed in FEV1 

responses between square-wave and triangular exposure profiles (Adams, 2003, 2006; Hazucha 

et al., 1992). These observations suggest the potential for somewhat enhanced O3 effects on 

lung function in ambient exposure scenarios that typically involve gradually increasing 

daily O3 exposure up to a peak in late afternoon and a subsequent gradual decline (i.e., the 

“triangular” exposure paradigm). 

The development of effects is time-dependent during both exposure and recovery periods, 

with considerable overlap of evolving and receding effects. In healthy human subjects exposed 

to typical ambient O3 concentrations (i.e., <0.12 ppm), spirometric responses largely resolve 

within a few hours (4 to 6 h) postexposure, but cellular effects persist for longer periods (~24 h). 

In such subjects, persisting small residual lung function effects are almost completely resolved 

within 24 hours; but in hyperresponsive individuals, the recovery takes longer (as much as 48 h) 

to return to baseline values. The majority of these responses are attenuated after repeated 

exposure, but such tolerance to O3 is lost within one week postexposure. Unfortunately, no new 

data are available on pulmonary function changes in animals upon chronic exposure to O3; but, 

8-19
earlier research on repeated exposure of rats to an episodic profile of O3 found small (but 

significant) decrements in lung function that were consistent with early indicators of focal 

fibrinogenesis in the proximal alveolar region. 

In the 1996 O3 AQCD, O3-induced decreases in inspiratory capacity were hypothesized to 

result from neurogenic inhibition of maximal inspiration due to stimulation of C-fiber afferents, 

either directly or from O3-induced inflammatory mediators. Earlier human studies (Coleridge 

et al., 1993; Hazucha and Sant’Ambrogio, 1993) reported a role for bronchial C-fibers and 

rapidly adapting receptors as primary vagal afferents responsible for O3-induced changes in 

ventilatory rate and depth. As discussed in Chapter 6, the newer results of Passannante et al. 

(1998) also support C-fiber stimulation as a primary mechanism of the O3-induced reduction in 

inspiratory capacity and suggest a role for nociceptive mechanisms. This neurogenic mechanism 

is also likely related to effects such as increased airway responsiveness and lung inflammation. 

Only two studies evaluated lung function changes due to O3 exposure in patients with 

preexisting respiratory diseases under experimental controlled exposure regimens. Studies using 

COPD patients found that O3 exposure-induced minimal effects (nonsignificant decreases in lung 

function) in this population (Gong et al., 1997a). However, newer studies of asthmatics by 

Alexis et al. (2000) still continue to indicate that asthmatics are at least as sensitive, if not more, 

than healthy subjects, based on pulmonary function deficiencies detected by spirometric 

analyses. 

In addition to effects of O3 exposure on the large airways as indicated by spirometric 

responses, O3 exposure also affects the function of the small airways and parenchymal lung 

tissue. Studies reported by Foster et al. (1993, 1997) that examined the effect of O3 on 

ventilation distribution in healthy adult males suggest a prolonged O3 effect on the small airways 

and ventilation distribution in some individuals. Animal toxicology studies have shown the 

centriacinar region (CAR) of the lung (the segment between the last conducting airway and the 

gas exchange region) to be a region highly susceptible to O3-induced damage (epithelial cell 

necrosis and remodeling of respiratory bronchioles). Common pulmonary function tests do not 

measure acute changes in the small airways of the CAR. Identification of acute effects of O3 in 

small airways, if any, would lend additional support for concerns about long-term effects of 

repeated O3 exposures. 

8-20
Airway Responsiveness 

The 1996 O3 AQCD recognized the induction of airway hyperresponsiveness (AHR) in 

humans on exposure to O3, which usually resolves in 18 to 24 h after exposure in a majority of 

subjects, but may persist in some individuals for longer periods. Ozone-induced nonspecific 

AHR could have clinical implications in asthmatics, possibly putting them at potential increased 

risk for more prolonged bouts of bronchoconstriction in response to various triggering stimuli in 

the ambient air. Three new studies (discussed in Chapter 6) suggest that (1) subjects with 

asthma developed tolerance to repeated O3 exposures in a manner similar to normal subjects; but 

there were more persistent effects of O3 on airway responsiveness, which only partially 

attenuated when compared to filtered air control exposures (Gong et al., 1997b); (2) the 

enhancement of allergen responsiveness after O3 exposure appears to be time-dependent, 

suggesting that the timing of allergen challenge in O3-exposed subjects with allergic asthma is 

important (Jörres et al., 1996); and (3) subjects with rhinitis exhibit significant, clinically 

relevant decreases in pulmonary function in the early phase allergen response (Holz et al., 2002). 

These observations suggest that O3 exposure may be a clinically important factor that can 

exacerbate responses to other ambiently-encountered bronchoconstrictor substances in 

individuals with preexisting allergic asthma and that its influence may be both immediate and 

persist for relatively long periods of time (a few days). 

An extensive laboratory animal study database (see Chapter 5 Annex Tables) exploring 

the effects of acute, long-term, and repeated exposures to O3, clearly indicates that induction of 

AHR occurs at relatively high O3 concentrations. Only one study that utilized very low O3 

(0.05 ppm) concentrations found that acute exposure induced AHR in a few inbred strains of rat, 

suggesting a role for genetic susceptibility in this process (Depudyt et al., 1999). As seen with 

humans, acute changes in AHR do not persist upon long-term exposure of laboratory animals to 

near-ambient concentrations of O3; and attenuation has been observed. Both human and animal 

studies indicate that airway responses are not associated with inflammation, but they do suggest 

a likely role for neuronal involvement. 

Lung Inflammation, Permeability, and Biochemical Alterations 

Respiratory tract inflammation and increased cellular permeability are the two most 

important biological markers of O3-induced injury response mechanisms in both humans and 

8-21
animals. These distinct, independent biological effects have been observed in all species studied 

in response to acute O3 exposure. 

Several controlled human exposure studies reviewed in the 1996 O3 AQCD clearly showed 

that a single acute exposure of humans to moderate O3 concentrations ($0.08 to 0.1 ppm) while 

exercising at moderate to heavy levels disrupts the barrier function of the lung, initiating a 

cascade of responses indicative of increased lung permeability and pulmonary inflammation, 

as indexed by a variety of cellular and biochemical changes, e.g., increases in levels of 

polymorphic neutrophils (PMNs) and protein in lung fluid. Both the inflammatory response and 

increased lung permeability have been observed as early as 1 h and persisted for at least 18 h. 

The newer studies reviewed in this document (see Chapter 6) provide additional information on 

three different aspects of O3-induced inflammatory responses, such as (1) intersubject variability; 

(2) differential attenuation profiles for different inflammatory markers; and (3) effects of 

repeated exposures. 

Soluble mediators of inflammation (e.g., the cytokines IL-6 and IL-8), as well as 

arachidonic acid metabolites (e.g., PGE2, PGF2", thromboxane, and leukotrienes [LTs] such 

as LTB4), have been measured in the BAL fluid of humans exposed to O3. In addition to their 

role in inflammation, many of these compounds have bronchoconstrictive properties and may be 

involved in increased airway responsiveness following O3 exposure. The time course for the 

inflammatory responses (including recruitment of neutrophils and other soluble mediators) is not 

clearly established, but differential attenuation profiles for many of these parameters are evident 

from the meta-analysis by Mudway and Kelly (2004) of 21 controlled human exposure studies. 

Recent studies continue to support the observation made in the 1996 O3 AQCD that 

repeated O3 exposures in humans also induce ongoing cellular damage irrespective of attenuation 

of the inflammatory responses and lung function decrements (Devlin et al., 1997; Jörres et al., 

2000). Devlin et al. (1997) studied inflammatory responses of humans repeatedly exposed to 

0.4 ppm O3 for 5 consecutive days. Several indicators of inflammation (e.g., PMN influx, IL-6, 

PGE2, fibronectin) were attenuated after 5 days of exposure (i.e., values were not different from 

FA). Several other markers (LDH, IL-8, total protein, epithelial cells) did not show attenuation, 

indicating that tissue damage probably continues to occur during repeated exposure. Recovery 

of the inflammatory response occurred for some markers after 10 days, but some did not return 

to normal even after 20 days. When re-exposed 2 weeks later, changes in BALF indicated that 

8-22
epithelial cells appeared to be fully repaired (Devlin et al., 1997). Kopp et al. (1999) observed 

inflammatory responses only after the first O3 peak in summer; and their absence late in summer 

(even after exposure to higher levels of O3) may be due to attenuation of such responses upon 

repeated O3 exposures. 

Numerous studies reported acute O3-induced changes in lung epithelial permeability 

assessed by indirect assay (increased levels of albumin and protein in BALF). A few other 

studies demonstrated O3-induced epithelial cell permeability through direct assessment of 

clearance of 99mTc-DTPA (technetium-99m labeled diethylene triamine pentaacetic acid). 

For example, Kehrl et al. (1987) showed increased 99mTc-DTPA clearance in healthy young 

adults at 75 minutes postexposure to 0.4 ppm O3 for 2 h. More recently, Foster and Stetkiewicz 

(1996) have shown that increased 99mTc-DTPA clearance persists for at least 18 to 20 h post-O3 

exposure (130 min to average O3 concentration of 0.24 ppm), and the effect is greater at the lung 

apices than at the base. 

Laboratory animals, like humans, exhibit varying degrees of sensitivity to O3 exposure 

(see Chapter 5 for detailed discussion); and this is evident even for the induction of pulmonary 

inflammation and permeability. New animal toxicology studies of O3-induced inflammation 

assessed in Chapter 5 indicate that the lowest acute O3 exposures that had an effect on mouse 

lung inflammation was 0.12 ppm for 24 hours. Shorter durations (8 h) required greater O3 

exposure (0.26 ppm) for effects on epithelial permeability but had no effect on inflammation. 

The lowest acute O3 concentrations that had an effect on epithelial permeability or inflammation 

in the rat were 0.5 ppm for 3 hours or 0.12 ppm for 6 hours. Also, increased lung permeability 

and inflammation occurred in rabbits with O3 exposures as low as 0.10 ppm for 2 h/day for 

6 days. Subchronic exposures in animals suggest that permeability changes are transient (and 

species-dependent) and return to control levels even with continuing exposure. Chronic animal 

O3 exposure studies suggest a role for persistent inflammation in O3-induced alterations in lung 

structure and function. Significant remodeling of epithelium and underlying connective tissues 

in distal airways have been reported in rats exposed to 0.25 ppm O3 (12 h/day for 6 wk) and in 

monkeys exposed to 0.2 ppm O3 (8 h/day for 90 d). Various factors such as viral infection, 

chemotactants and oxidized matrix fragments are also implicated in the establishment and 

persistence of O3-induced inflammation. 

8-23
Animal toxicology and human in vitro studies that evaluated biochemical mediators 

implicated in injury and inflammation found alterations in the expression of cytokines, 

chemokines, and adhesion molecules, indicative of an ongoing active stress response as well as 

injury repair and regeneration processes. Both animal and human studies indicate cellular and 

biochemical changes associated with inflammation and increased permeability, but the 

relationship between these changes and possible airway remodeling is not known. 

As discussed in the earlier section on dosimetry, interaction of O3 with the lipid 

constituents of pulmonary surfactant has been proposed as one of the key mechanisms by which 

O3 exerts its toxic effects. Experimental evidence clearly indicates a role for the interaction of 

O3 with lipid constituents of the ELF and cell membranes and the generation of lipid ozonation 

products and secondary redox mediators in the initiation of site-specific cell-injury response 

cascades. One such lipid ozonation product, 4-hydroxynonenal, has been found to bind to 

proteins and increased protein adducts in human alveolar macrophages, suggesting a role for 

4-hydroxynonenal in acute cell toxicity. Cholesterol, the most abundant neutral lipid in 

pulmonary surfactant is susceptible to attack by O3, resulting in formation of multiple oxidized 

cholesterol products, e.g., cholesterol epoxide. A 20-fold increase in cholesterol epoxide in the 

BALF from mice exposed to 0.5 ppm O3 for 3 h suggests a potential role for this oxidation 

product in O3 toxicity (Pulfer et al., 2005). Species- and region-specific increases in lung 

xenobiotic metabolism have been observed in response to both short- and long-term O3 

exposure. It has been well recognized that antioxidants in the ELF confer some protection 

against O3 toxicity. But even with environmentally relevant exposures, O3 reactivity is not 

quenched. Further, antioxidant reactivity with O3 is both species-specific and dose-dependent. 

Lung Injury 

Pulmonary histopathological observations reported in the 1996 O3 AQCD suggested that 

similar types of alterations occurred in lung morphology in all laboratory animal species studied, 

including primates, upon short-term O3 exposure. The cells in the CAR were recognized as a 

primary target, possibly because the CAR receives the greatest dose of O3 delivered to the lower 

respiratory tract. With chronic O3 exposure, structural changes were also observed in this region 

of the respiratory tract (the region typically found affected in most chronic airway diseases of the 

human lung). Simulated seasonal exposure studies in animals also suggested that seasonal O3 

8-24
exposures may have potential for cumulative impact over many years. Ciliated cells in the nasal 

cavity and airways and Type I epithelial cells in the gas-exchange region have also been 

identified as targets. Though acute O3 exposure induces structural changes such as fibrosis in the 

CAR, these structural alterations appear to be partially transient, with recovery shortly post O3 

exposure; but the time for recovery is dependent on the species and the dose of O3. Long-term or 

prolonged exposure to O3 has been found to cause chronic lesions similar to early lesions of 

respiratory bronchiolitis, which have the potential to progress to fibrotic lung disease. Some of 

the morphological changes associated with long-term O3 exposures, e.g., increases in 

hyperplastic epithelial cells, appear to reverse following cessation of O3 exposure. However, 

in the underlying interstitium of the CAR, proliferation of fibroblasts creates excess noncellular 

matrices. These processes are only partially reversible and may progress following cessation of 

O3 exposure. This suggests initiation of focal interstitial fibrosis, which can progress to 

nonreversible structural damage to lung tissue. 

Reports of morphological changes following chronic O3 exposures in animal studies 

(rodents and primates) published since the 1996 AQCD allude to the earlier findings assessed in 

that document. In rats, the effects of chronic ~0.5 ppm O3 exposure included mucous cell 

metaplasia, hyperplasia of the nasal epithelium, increased mucosubstances, and increased Bcl-2 

protein levels. In mice, lifetime exposures of 0.5 ppm O3 were linked to similar outcomes. 

Taken together, the rodent studies suggest that O3 exposure may have the potential to induce 

similar long-lasting alterations in human airways. A series of new studies that utilized infant 

rhesus monkeys and simulated seasonal ambient exposure (0.5 ppm 8 h/day for 5 days, every 

14 days for 11 episodes) reported remodeling in the distal airways; abnormalities in tracheal 

basement membrane; eosinophil accumulation in conducting airways; and decrements in airway 

innervation, again confirming the potential greater injury due to seasonal exposure compared to 

continuous exposure alluded to in the 1996 O3 AQCD. 

Host Defense 

Based on a small number of studies available at the time, the 1996 O3 AQCD concluded 

that short-term O3 exposure of laboratory animals and humans impairs alveolar macrophage 

(AM) clearance of viable and nonviable particles from the lungs and decreases the effectiveness 

of host defenses against bacterial lung infections in animals and perhaps humans. A single 

8-25
controlled human exposure study reviewed in the 1996 O3 AQCD found decrements in the 

ability of alveolar macrophages to phagocytose microorganisms upon exposure to 0.08 to 

0.1 ppm O3 for 6.6 h during moderate exercise (Devlin et al., 1991). 

Other evidence for O3-induced dysfunction of host defense components and for subsequent 

enhanced susceptibility to bacterial lung infection has been derived from laboratory animal 

studies. Acute exposures of 0.08 ppm (3 h) O3 were found to result in the mortality of mice due 

to Streptococcal bacterial infection. Changes in antibacterial defenses appear to be dependent on 

exposure regimens, species and strain of test animal, species of bacteria, and age of animal (with 

young mice being more susceptible to the effects of O3, for example). Animal toxicology studies 

indicated that acute O3-induced suppression of alveolar phagocytosis and immune functions 

observed in animals appeared to be transient and attenuated with continuous or repeated 

exposures. 

It has also been reported that O3 exposures can interfere with AM-mediated clearance in 

the respiratory region of the lung and with mucociliary clearance of the tracheobronchial 

airways. Ozone-induced perturbations in the clearance process have been found to be dosedependent, with low dose exposures accelerating clearance and high doses slowing the clearance 

process. Some respiratory tract regional- and species-specific differences have also been 

observed. 

In vitro cultures of epithelial cells obtained from nonatopic and mild atopic asthmatics 

exposed to 0.01 to 0.1 ppm O3, exhibited significantly increased permeability compared to cells 

from normal subjects, thus indicating a potential inherent susceptibility of cells from asthmatics 

for O3-induced permeability. Newer in vitro cell culture studies of human bronchial epithelial 

cells indicate O3-induced exacerbation of human rhinovirus type 16 infection (Spannhake et al., 

2002); and new animal toxicology studies have shown O3-induced modulation of cell-mediated 

immune responses affecting the onset and persistence of infection in rats (Cohen et al., 2001, 

2002). 

The available data at this time indicate that acute O3 exposure has a potential to impair host 

defense capability, primarily by interfering with the functions of alveolar macrophages. Any 

impairment in macrophage function may lead to decreased clearance of microorganisms or 

nonviable particles. Compromised alveolar macrophage functions in asthmatics may increase 

their susceptibility to other O3 effects, the effects of particles, and respiratory infections. 

8-26
Cardiovascular Effects 

Based on the paucity of then-available information, the 1996 O3 AQCD did not accord 

much, if any, attention to possible O3-related cardiovascular effects. However, since then, 

an emerging body of animal toxicology evidence is beginning to suggest that hematological and 

thermoregulatory alterations (in heart rate and/or core body temperature) may mediate acute O3 

cardiovascular effects. For example, it is beginning to be recognized that O3-induced lung injury 

and permeability changes, as well as O3-induced alterations in hemodynamics, may lead to 

cardiovascular system effects. Also, interactions of O3 with ELF lipids and surfactants result in 

lipid ozonation products and reactive oxidant species (ROS) that have the potential to penetrate 

the epithelial barrier and to initiate toxic effects on the cardiovascular system. 

Earlier studies in rats indicated a potential role for platelet activating factor (PAF) in the 

O3-induced inflammatory response. Recent observations of O3-induced generation of oxysterols 

and $-epoxides from cholesterol in surfactant also suggest that these lipid ozonation products 

(like lysophospholipids) might exhibit PAF-like activity and contribute to clotting and 

thrombolytic effects in the cardiovascular system. 

Other studies carried out using isolated perfused rat lung model (Delaunois et al., 1998) 

indicate inhibition of pulmonary mechanical reactivity to bronchoconstrictors and persistent 

vasoreactivity of the vascular bed upon exposure to O3 (0.4 ppm for 4 h). Newer studies have 

also now found that acute (less than 5 h) 0.1 ppm O3 exposure caused decreased heart rate in 

young, but not old, rats (Arito et al., 1997). 

Only one human experimental O3 exposure study (Gong et al., 1998) evaluated potential 

cardiovascular effects in normal and hypertensive adult males. Various cardiovascular and 

hemodynamic parameters were monitored upon exposure to O3 (0.3 ppm for 3 h) with 

intermittent exercise. No significant O3-induced differences were observed in ECG, heart rate, 

or blood pressure in either normal or hypertensive subjects. An overall increase in myocardial 

work and impairment in pulmonary gas exchange was observed, however, that might be 

clinically important in patients with preexisting cardiovascular impairment with or without 

concomitant lung disease. 

8-27
8.4.3 Biological Basis for O3 Health Effects Assessment 

The scientific knowledge base gained from animal toxicological studies and experimental 

studies of human volunteers in clinical settings (discussed in Chapters 4, 5, and 6 of this 

document) has notably expanded the knowledge base beyond that available at the time of 1996 

O3 AQCD. This section provides an interpretive integration of key findings derived from the 

experimental knowledge base that can aid in evaluating the biological plausibility for health 

effects observed in O3-related epidemiological studies discussed later in this chapter. 

Animal-to-Human Extrapolation Issues 

The physiological and biochemical observations summarized in Table 8-1 provide an 

overview of the main types of acute O3-induced health effects as demonstrated by toxicological 

studies of humans and animals. This table was generated from those experimental studies (see 

Annexes for Chapters 5 and 6 for experimental details) that utilized exposure regimens of varied 

concentration and duration that are environmentally relevant. As noted above in Section 8.4.1, 

observed acute O3 effects are mostly transient (generally persisting #24 to 48 h) and attenuate 

over time. However, the time-line for resolution of many of these physiological and biochemical 

parameters in normal and human subjects with underlying cardiopulmonary diseases follow 

different profiles, as presented in Figure 8-3. Alterations in the cellular and molecular profiles 

observed in human airway epithelium upon acute exposure to O3 evolve over time (Figure 8-3), 

and the knowledge of this profile is valuable in assessing biological plausibility to integrate 

across evidence for various health endpoints. 

Basic similarities in physiological, biochemical, and pathological processes that exist 

between human and other mammalian species are derived from the high degree of genome 

sequence homology that exists across various species. This homology reinforces the 

significance of knowledge gained on the initiation, progression and treatment regimes for 

various disease processes across animal species. This homology is also apparent in acute O3­

induced effects, especially on the respiratory tract of human and animal species as presented in 

Table 8-1 and Figures 8-3 and 8-4. The commonality of phenomenon observed in humans and 

rats with regard to respiratory system effects (in terms of spirometry, ventilatory response, host 

defense and inflammation) and their attenuation adds strength to animal-human extrapolations. 

Such similarities observed at higher levels of cellular organization (neutrophilic inflammation, 

8-28
Table 8-1. Acute O3-Induced Physiological and Biochemical 

Changes in Human and Animals 

Physiological/Biochemical 

Alterations Human Exposure Studies 1,2 Animal Toxicology Studies 3,4 

Pulmonary Function: 9 FEV1 

8 Frequency of breathing 

(rapid, shallow ) 

9 FVC 

8 Frequency of breathing

 (rapid, shallow ) 

9 FVC 

(cough, breathing discomfort, 

throat irritation, wheezing) 

Mild bronchoconstriction 

Airway Responsiveness: 8 (neuronal involvement) 

Change in lung resistance 

8 (vagal mediation) 

Change in lung resistance 

Inflammation: Yes Yes 

8 inflammatory mediators 8 inflammatory mediators 

Reactive Oxygen Species: 8 8 

Host Defense: 8 particle clearance 8 particle clearance 

8 permeability 8 permeability 

9 AM phagocytosis 9 clearance of bacteria 

8 severity of infection 

8 mortality & morbidity 

Lung Injury: 

Morphology: Yes Yes 

Susceptibility: Age, 

Interindividual variability 

Disease status 

Species-specific differences 

Genetic basis for susceptibility 

indicated 

Polymorphism in certain genes 

being recognized 

Cardiovascular Changes: Impairment in arterial O2 transfer 

Ventilation-perfusion mismatch 

(suggesting potential arterial 

vasoconstriction) 

8 rate pressure product5 

8 myocardial work5 

Heart rate 

9 core body temperature 

8 atrial natriuretic factor 

Role for platelet activity factor 

(PAF) indicated 

Increased pulmonary vascular 

resistance 

1

 Controlled chamber exposure studies in human volunteers were carried out for a duration of 1 to 6.6 h with O3

 concentration in the range of 0.04-0.40 ppm with intermittent exercise. See text for discussion of O3 levels

 shown to cause different types of effects listed.

2

 Data on some biochemical parameters obtained from in vitro studies of cells recovered from BALF. 3

 Responses were observed in animal toxicology studies with exposure for a duration of 2 to 72 h with O3

 concentration in the range of 0.1 to 2.0 ppm. 

4

 Various species (mice, rat, guinea pigs and rabbit) and strains. 

5

 In hypertensive subjects. 

8-29
8-30 

Figure 8-3. Resolution time-line for the respiratory, physiological, and biochemical parameters are d

from studies reported in Chapter 6 and Chapter 6 Annex. 
Figure 8-4. Acute (1-8 h) O3 exposure-induced cellular and molecular changes and 

timelines for their resolution depicted here are derived from the data 

reported in Leikauf et al. (1995) and Mudway and Kelly (2000). 

macrophage phagocytosis processes) have increased the value and importance of animal studies 

in generating important data that are impossible to collect in human studies but which may 

corroborate both human clinical and epidemiologic studies. 

Extrapolation of results derived from laboratory animal studies to humans involves a 

combination of dosimetry, end point homology and species sensitivity, particularly in the case of 

exposure and health outcome analyses. However, existing extrapolation models have not yet 

been sufficiently validated to allow for highly confident quantitative animal-to-human 

extrapolation for O3 effects. Still, qualitative extrapolation appears to be reasonable for some 

endpoints. For example, based on inflammatory markers in BALF, a 2 ppm O3 exposure in 

nonexercising rats approximates to a 0.4 ppm exposure in exercising humans (Hatch et al., 

1994). This observation lends support to the use of some of the animal toxicology data derived 

from relatively high O3 concentration exposure regimens in understanding putative molecular 

changes likely to be associated with acute O3 exposure in humans. 

8-31
The time courses for induction and resolution of various types of O3 effects experimentally 

induced in laboratory settings (see Figures 8-3 and 8-4) may provide a basis for reasonable 

projection of likely lag times for observing associations between ambient O3 concentrations/ 

exposures and analogous effects in epidemiologic studies of human populations. For example, 

it would not be unreasonable to expect that pulmonary function decrements could be detected 

epidemiologically within lags of 0 (same day) or 1 or 2 days following O3 exposure, given the 

rapid onset of spirometric changes and their persistence for 24 to 48 hours among 

hyperresponsive human subjects in clinical studies. On the other hand, although asthmatic 

individuals may begin to experience symptoms soon after O3 exposure, it may take a day or so, 

coupled with breathing difficulties secondary to O3-induced hyperresponsiveness that may 

persist for 2 to 3 days, for members of that sensitive population group to seek medical attention. 

This could possibly be reflected by epidemiologic observations of significantly increased risk for 

asthma-related ED visits or hospital admissions with 1- to 3-day lags, or, perhaps, enhanced 

distributed lag risks (combined across 3 days) for such morbidity indicators. Analogously, one 

might project increased mortality within 0 to 3 day lags as a possible consequence of O3-induced 

increases in prothrombotic agents, such as platelet aggregating factor (PAF) or lipid ozonation 

products (e.g., cholesterol epoxide) arising from inflammation cascades occurring within 12 to 

24 hours of O3 exposure (see Figure 8-4). 

Similarly, the presence of apparent O3-induced lesions in animals from chronic O3 

exposure studies (12 to 24 months) indicate morphological alterations that may analogously 

occur in humans with long-term (months, years) chronic exposure to relatively high O3 levels, 

but specific O3 concentrations and exposure patterns that may produce analogous alterations in 

human lungs remain to be substantiated. 

8.4.4 Epidemiologic Evidence 

Epidemiologic evidence available at the time of the 1996 O3 AQCD indicated potential 

health effects associated with acute ambient O3 exposures. The 1996 O3 AQCD further stated 

that only suggestive epidemiologic evidence existed for health effects of chronic ambient O3 

exposure in the population, and this was partly due to an inability to isolate potential effects 

related to O3 from those of other pollutants, especially PM (U.S. Environmental Protection 

Agency, 1996a). The availability of numerous recent epidemiologic studies, with some studies 

8-32
designed explicitly to evaluate ambient O3 exposure effects (as discussed in detail in Chapter 7 

of this document), makes it possible to better assess the health effects of both acute and chronic 

ambient O3 exposure. There exist a number of issues and uncertainties associated with the 

interpretation of O3 health effects evidence in epidemiologic studies, which are discussed in 

Sections 7.1.3 and 7.6 of Chapter 7 in this document. These include the use of various indices to 

represent O3 exposure, exposure measurement errors (i.e., differences between ambient 

concentrations and personal exposure), lag period between exposure and effect, potential 

confounding by temporal and meteorological factors, potential confounding by copollutants, the 

concentration-response function, and heterogeneity of O3 health effects. 

The recent epidemiological studies that have been conducted in areas across the United 

States and Canada, as well as in Europe, Latin America, Australia and Asia, are reviewed in 

Chapter 7 (for details see Annex to Chapter 7). Present discussions in this integrative synthesis 

chapter focus on findings from studies carried out in the United States and Canada. Important 

features of these epidemiologic studies (e.g., study location, study population, mean O3 

concentrations, health risk estimates) are summarized in Chapter 8 Appendix Tables 8A-2 to 

8A-5. The studies are ordered in those tables by the 98th and 99th percentile values of the 

calculated 8-h max O3 data for the full study period. The 98th and 99th percentile values were 

selected as they represent a high concentration that roughly approximates a 4th maximum 

concentration, depending on the study period length. For studies that did not have this data 

available, their ordering was approximated based on the mean O3 concentrations observed in 

the study. 

8.4.4.1 Acute Ozone Exposure Studies 

Numerous epidemiological studies carried out over the past decade have added evidence to 

the knowledge base assessed in the 1996 O3 AQCD, which included both (a) individual-level 

summer camp and exercise studies that established a relationship between ambient O3 exposure 

and lung function decline and (b) aggregate population studies that suggested positive 

relationships for O3-related respiratory morbidity endpoints (i.e., respiratory ED visits and 

hospitalizations). The new studies reviewed in Chapter 7 in this document include numerous 

field/panel studies and population studies from various regions in the United States and abroad. 

In field/panel studies on the effects of air pollution exposure, the most common health outcomes 

8-33
measured were lung function and respiratory symptoms. The population studies examined daily 

ED visits, hospital admissions, and mortality data. 

Field/Panel Studies of Acute Exposure Effects 

Pulmonary Function and Respiratory Symptoms 

Many of the new field/panel studies reviewed in Chapter 7 and the controlled human 

exposure studies reviewed in Chapter 6 of this document provide additional data supporting two 

key findings reported in the 1996 O3 AQCD. One key finding was that acute O3 exposure was 

associated with a significant decline in lung function parameters. Ozone-related lung function 

decrements were most notable in children, as indicated by the results from the meta-analysis of 

summer camp studies by Kinney et al. (1996). Similar responses were reported for children and 

adolescents exposed to O3 in ambient air in the summer camp studies as were found with 

exposure to O3 in clean air for 1 to 2 h while exercising in clinical studies. A second key finding 

was that adults who work or exercise outdoors were also found to be vulnerable to O3-associated 

declines in lung function due to their greater exposure to O3 during periods of increased physical 

activity. 

In a number of newly available field/panel studies, FEV1 was measured in panels of 

exercising children, outdoor workers, and adult hikers exposed to ambient O3 while experiencing 

elevated exertion levels. Collectively, the results of the new studies (discussed in 

Section 7.2.3.1) confirm and extend the findings from analogous field/panel studies and 

experimental controlled human exposure studies assessed in the 1996 O3 AQCD. Acute O3 

exposures prolonged over several hours and combined with elevated levels of exertion or 

exercise were found to have magnified effects on lung function, as evaluated in terms of FEV1. 

Only studies carried out in the U.S. and Canada that evaluated lung function changes are 

discussed here (see Appendix Table 8A-2 for details). 

Brauer et al. (1996) measured lung function in berry pickers in British Columbia, Canada. 

The mean ambient 1-h max O3 was 40.3 ppb (SD 15.2). Significant O3-related decreases in FEV1 

were observed in the morning and afternoon, but not across the day. Another field study on 

adult hikers by Korrick et al. (1998) observed that the mean FEV1 decline in healthy hikers was 

marginal, but a 4-fold greater decline was observed in hikers with asthma and wheeze. Both 

studies, however, did not evaluate the possible role of confounding by copollutants. Other 

8-34
studies from Germany and Mexico City provided additional supportive evidence that healthy 

adults with prolonged periods of exposure and exertion experienced declines in lung function. 

Several panel studies that examined lung function changes in children, including a few 

U.S. studies (Neas et al., 1995, 1999; Linn et al., 1996), observed small declines in FEV1 or in 

peak expiratory flow (PEF). It should be noted, however, that the group mean responses 

reported in panel or population studies may mask interindividual differences, e.g., larger lung 

function changes in sensitive populations (as discussed in Section 8.4.2 and in more detail in 

Chapter 6). 

Panel studies that evaluated lung function decrements in asthmatic children collectively 

indicated O3-induced decrements, although most of the individual-study estimates were not 

statistically significant. Results from a multicity study of asthmatic children from eight urban 

areas in the U.S. (Mortimer et al., 2002) suggest a small, but significant decline in PEF based on 

a multiday distributed lag model. Of all the pollutants examined, including O3, PM10, NO2, and 

SO2, only O3 was found to be associated with morning PEF. The small change in lung function 

observed likely reflects the low mean O3 levels across the eight cities studied (8-h avg O3 of 

48 ppb). Overall, acute ambient O3 exposure was observed to be associated with lung function 

decrements at O3 concentrations ranging down to a 98th percentile 8-h max of 55 to 60 ppb for 

potentially vulnerable and susceptible populations, as shown in Appendix Table 8A-2. 

In addition to pulmonary function, the majority of asthma panel studies also evaluated the 

relationship between O3 and respiratory symptoms in asthmatic children. The results obtained 

from these studies show some inconsistencies, with some indicating significant positive 

associations and other well-conducted, albeit relatively smaller, studies not finding such effects. 

Overall, however, the multicity study by Mortimer et al. (2002) and several credible single-city 

studies (e.g., by Gent et al., 2003) indicate a fairly robust positive association between ambient 

O3 concentrations and increased respiratory symptoms in asthmatics. 

Airway Inflammation 

A few epidemiologic studies, conducted in Germany and Mexico City, examined the effect 

of acute ambient O3 exposure on airway inflammation in children. Results indicated that 

associations were found largely on days or in study locations having the higher O3 

concentrations (1-h max O3 of approximately 100 ppb) among those present. At locations with 

8-35
lower O3 levels, no association was detected between upper airway inflammation and ambient O3 

concentrations across the entire study period (Kopp et al., 1999). However, a more detailed 

analysis showed that the first significant O3 episode of the summer was associated with increases 

in inflammatory markers, but subsequent and even higher O3 episodes had no effect, suggesting 

an adaptive response. 

Cardiophysiological Endpoints 

A few recent air pollution epidemiology studies have also evaluated the potential effects of 

ambient O3 on cardiophysiological endpoints. Based on the discussions presented in Chapter 7 

(see Section 7.2.7), the limited data available at this time are suggestive of altered heart rate 

variability (HRV), ventricular arrhythmias, and MI incidence possibly being associated with 

acute O3 exposures; but, overall, these findings must currently be considered inconclusive. 

Population Studies of Acute Exposure Effects 

Emergency Department Visits and Hospital Admissions 

Many time-series studies reviewed in the 1996 O3 AQCD indicated positive associations 

between O3 air pollution and increased respiratory hospital admissions and ED visits, providing 

strong evidence for a relationship between O3 exposure and increased exacerbations of 

preexisting respiratory disease in the general public at 1 h-max O3 concentrations <0.12 ppm 

(120 ppb). Analyses of data for the northeastern United States suggested that O3 air pollution 

was associated with a substantial portion (on the order of 10 to 20%) of all summertime 

respiratory hospital visits and admissions. Several new studies have been published in the past 

decade examining temporal associations between O3 exposures and emergency department visits 

and hospital admissions for respiratory diseases. 

Studies conducted in the United States and Canada examining the association between 

acute O3 exposure and asthma ED visits are summarized in Appendix Table 8A-3. The risk 

estimates presented in Appendix Table 8A-3 include those using all-year and warm-season only 

O3 data. A majority of the studies conducted in the U.S. and Canada examined the effects of O3 

during the warm season only. An analysis of respiratory disease ED visits in Atlanta, GA over 

an 8-year period indicated a significant positive association between ambient O3 concentrations 

and asthma visits during the warm season (Peel et al., 2005). The mean 8-h max O3 from March 

8-36
to November was 55.6 ppb (SD 23.8) in this study. A study conducted in Saint John, Canada 

observed a positive association between O3 and asthma emergency visits during the warm season 

in individuals 15 year or older at even lower O3 concentrations (mean 1-h max O3 of 41.6 ppb) 

(Stieb et al., 1996). However, several other studies, mostly using all-year data, reported no 

association between O3 and ED visits for respiratory causes. In general, ambient O3 

concentrations were associated with ED visits for asthma in warm-season only analyses. 

Relationships between ambient O3 and respiratory-related hospital admissions have also 

been evaluated in various epidemiologic studies carried out in the United States and abroad. 

Results from studies in the U.S. and Canada are summarized in Appendix Table 8A-4. Some of 

these studies evaluated potential confounding by season or copollutants of the O3 effects on 

respiratory hospitalizations. A multicity study by Burnett et al. (1997a) evaluated the 

relationship between ambient O3 and respiratory hospitalizations in 16 Canadian cities over an 

11-year period. Seasonal analyses indicated that O3 effects were only observed during the warm 

season. As shown in Appendix Table 8A-4, most studies in the United States and Canada 

indicated a consistent positive association between ambient O3 concentrations and respiratory 

hospital admissions in the warm season, including studies with 98th percentile 8-h max O3 levels 

as low as about 50 ppb. Analyses of all-year data were mostly positive, but nonsignificant for 

either total respiratory or asthma hospital admissions. The studies that found positive 

associations based on analyses using various lag periods found the association to be strong 

with short lag periods (0 to 2 day), suggesting likely short-term O3 effects on asthma hospital 

admissions. 

Some new epidemiologic studies have also evaluated the associations between ambient O3 

concentrations and cardiovascular-related hospitalizations (see Section 7.3.4 for details). The 

epidemiologic evidence for cardiovascular morbidity is much more mixed than for respiratory 

morbidity, with only one of several U.S./Canadian studies showing statistically significant 

positive associations of cardiovascular hospitalizations with warm-season O3 concentrations. 

Studies from Taiwan and Hong Kong also provided suggestive evidence of an association, but 

most of the available European and Australian studies (all of which conducted all-year O3 

analyses) did not find an association between short-term O3 concentrations and cardiovascular 

hospitalizations. 

8-37
Acute Effects of Ozone on Mortality 

The 1996 O3 AQCD noted that an association between O3 concentrations and daily 

mortality for areas with high O3 levels (e.g., Los Angeles) was suggested by a few studies, but 

the magnitude of such an effect was unclear. However, due to the limited number of available 

studies and uncertainties regarding weather model specifications, no conclusive evidence of O3­

mortality associations was found in the 1996 O3 AQCD. In contrast, newly available large 

multicity studies designed specifically to examine the effect of O3 on mortality have since 

provided much more credible information. Numerous recent air pollution epidemiologic studies 

have been published that examined acute O3 exposure effects on all-cause mortality in the U.S. 

and elsewhere. Studies from the U.S. and Canada are presented in Appendix Table 8A-5. 

Five new multicity studies, three conducted in the United States and two in Europe, found 

small, but very precise (extremely narrow 95% CIs) positive associations for increased mortality 

risk using all-year ambient O3 data and warm-season only data. Most notably, the largest of the 

multicity studies conducted in the U.S. that evaluated the association between ambient O3 

concentrations and excess mortality risk, the NMMAPS 95 U.S. communities study by Bell et al. 

(2004), observed a significant positive association between O3 and all-cause mortality using all 

available data (55 communities with all-year data and 45 communities with warm-season only 

data). When the data were restricted to the warm season only, a slightly smaller, but still 

significant, O3 effect was observed. Analyses suggested that the O3-related excess mortality risk 

estimate was not confounded by exposure to PM. Another multicity study by Schwartz (2005) 

examining the association between O3 and all-cause mortality observed a slightly greater excess 

risk with warm-season only data compared to all-year data using a case-crossover study design. 

Mortality risk estimates from single-city studies were also positive, but many were not 

statistically significant. However, as in the case of ED visits and hospitalizations, restricting 

data to the warm season resulted in mortality risk estimates that were more positive and stronger. 

Meta-analyses of single-city mortality studies (Bell et al., 2005; Ito et al., 2005; Levy et al., 

2005) observed that O3-mortality risk estimates from warm-season only analyses were 

approximately 2-fold greater than those from all-year analyses. As shown in Appendix Table 

8A-5, the recent multicity and single-city studies generally show consistent positive and 

significant associations between acute O3 exposure and all-cause mortality in studies with 98th 

percentile 8-h max O3 values of 80 to 85 ppb and above. 

8-38
Some studies examined specific subcategories of mortality; however, most of these studies 

had limited statistical power to detect associations due to the small daily counts for the specific 

causes of death. In studies using broad categories of death, such as cardiopulmonary mortality, 

significant associations with acute O3 exposure have been observed. For example, an analysis of 

the NMMAPS data by Huang et al. (2005) examined the association between ambient O3 

concentrations and cardiopulmonary mortality in 19 large U.S. cities during the warm season. 

Ozone was positively associated with excess risk in cardiopulmonary mortality, even after 

adjusting for PM and heat waves. Also, the meta-analysis by Bell et al. (2005) examined the 

excess risk estimates for cardiovascular and respiratory causes and observed a slightly greater 

risk of cardiovascular mortality compared to all-cause and respiratory disease mortality. It is 

notable that only one single-city study, from among approximately 30 studies (including 

2 multicity ones), did not find a positive association between ambient O3 concentrations and 

increased cause-specific cardiovascular mortality, with both the multicity and several single-city 

associations being statistically significant. It should be noted that these analyses of causespecific mortality do not provide direct evidence on possible causal mechanisms as they often do 

not incorporate information on the contributing causes of death. For example, if an individual 

has been suffering from a major cardiovascular disease, his/her death may be misclassified as 

cardiovascular, even if a respiratory condition causes the death. 

8.4.4.2 Chronic Ozone Exposure Studies 

There were only a limited number of studies reported in the 1996 O3 AQCD that addressed 

potential health effects of long-term ambient O3 exposures. The few available epidemiological 

studies that had attempted to associate chronic health effects in humans with long-term O3 

exposure provided very limited suggestive evidence that such a linkage may exist. Several 

longitudinal epidemiologic studies carried out during the past decade have further evaluated the 

potential effects of chronic (several weeks to many years) O3 exposure on lung function, 

respiratory symptoms, lung inflammation, asthma prevalence, and birth defects. The strongest 

evidence is for seasonal effects of extended O3 exposures on lung function in children, i.e., 

reduced lung function growth being associated with higher ambient O3 levels. Based on the 

available data at this time, however, no clear conclusions can be drawn regarding the 

relationship between chronic O3 exposure and the other health outcomes. 

8-39
Few studies have evaluated effects of long-term O3 exposure on the incidence of cancer 

and/or mortality. Uncertainties regarding possible exposure periods of relevance and, also, 

inconsistencies across mortality outcomes and gender raise concerns regarding plausibility. 

The largest and most representative U.S. study, by Pope et al. (2002), observed positive but 

nonsignificant associations between long-term O3 concentrations and all-cause, 

cardiopulmonary, and lung cancer mortality. Thus, there is currently little evidence for potential 

relationships between chronic O3 exposure and increased incidence of cancer and mortality. 

8.4.4.3 Summary of the Epidemiologic Evidence 

Assessment of the evidence from epidemiologic studies for various health outcomes has 

focused on several considerations that are important in forming judgments as to the likely causal 

significance of the observed associations. As discussed in Section 7.1.2, these include the 

strength of the epidemiologic evidence, including the magnitude and precision of reported O3 

effect estimates and their statistical significance, and the robustness of the effects associations, 

or stability in the effect estimates after considering a number of factors, including alternative 

models and model specifications, potential confounding by copollutants, as well as issues related 

to the consequences of measurement error. Consistency of the effects associations involves 

looking across the results of multiple- and single-city studies conducted by different 

investigators in different places and times. 

In general, when associations are strong in terms of yielding large relative risk estimates, 

it is less likely that the association could be completely accounted for by a potential confounder 

or some other source of bias. With associations that yield small relative risk estimates, it is 

especially important to consider potential confounding and other factors in assessing causality. 

Across the range of different health outcomes, effect estimates for many health associations 

reported with O3 are generally small in size and could thus be characterized as weak. For 

example, effect estimates for associations with mortality generally range from 0.5 to 5% 

increases per 40 ppb increase in 1-h max O3 or equivalent, whereas associations for 

hospitalization range up to about 50% increases per standardized O3 increment. Of particular 

note are several multicity studies that have yielded relative risk estimates for associations 

between short-term O3 exposure and mortality or morbidity that, although small in size, have 

great precision due to the statistical power of the studies. For example, corresponding summary 

8-40
estimates for mortality in large U.S. multicity studies ranged from 0.5 to 1.0% per standardized 

24-h increment, with some notable heterogeneity across cities. Thus, such associations are 

strong relative to the precision of the studies; that is, the associations were strong enough to have 

been reliably measured by the studies such that many of the associations can be distinguished 

from the null hypothesis with statistical confidence (e.g., at p # 0.05). 

PM

As discussed in Section 7.6.4.2, the associations reported between short-term O3 exposure 

and various health outcomes were generally robust to adjustment for copollutants, including 

2.5 and sulfates, which are correlated with O3 concentrations in many areas in the United 

States. More limited evidence is available on the effects of different modeling strategies on 

associations with O3; however, these studies indicate that O3 associations may be robust to 

various model specifications for temporal trend adjustment. Risk estimates for O3 were 

generally more sensitive to the use of alternative weather models than to adjustment for temporal 

trends, due to the temperature-dependent nature of O3 formation in the atmosphere. 

In considering results from the multicity studies and single-city studies in different 

geographic areas, there was general consistency in effects of short-term O3 exposure on 

mortality, respiratory hospitalization and other respiratory health outcomes. Some variation in 

effects was observed and that may be attributable to differences in relative personal exposure to 

O3, which is affected by factors such as air conditioning prevalence and activity patterns, as well 

as varying concentrations and composition of copollutants present in different regions. Thus, 

consideration of consistency or heterogeneity of effects is appropriately understood as an 

evaluation of the similarity or general concordance of results, rather than an expectation of 

finding quantitative results within a very narrow range. 

Taken together, the epidemiologic evidence shows strength, robustness, and consistency in 

associations between short-term O3 exposure and a range of respiratory morbidity health 

outcomes, including increased respiratory-related hospital admissions and asthma-related ED 

visits. There is also relatively strong epidemiologic evidence for associations between shortterm O3 exposure and all-cause mortality. There is less strong, but nevertheless highly 

suggestive evidence for effects related to cardiovascular morbidity. Moreover, consistently 

positive associations were found for O3-related cardiovascular mortality across approximately 

30 studies, with two well-conducted multicity studies in the United States and Europe yielding 

small but statistically significant positive associations. In contrast, generally inconclusive 

8-41
evidence currently exists for associations between long-term chronic O3 exposure and either 

morbidity or mortality. 

8.5 ASSESSMENT OF POTENTIAL THRESHOLDS 

A key issue in assessment of air pollution health effects is whether any thresholds can be 

discerned for various types of health effects. Animal toxicology studies reviewed in Chapter 5 

appear to indicate a linear concentration-response relationship for O3 exposure effects. Lowest 

concentrations found to have caused statistically significant health effect changes, e.g., altered 

pulmonary function, airway hyperresponsiveness, lung inflammation, increased lung tissue 

permeability, or altered host defense, in various laboratory animal species are listed in Appendix 

Table 8A-6. Some of these concentrations are rather high, but others appear to be more 

environmentally relevant, including at least one study where the induction of airway 

hyperresponsiveness was seen at O3 levels distinctly below current ambient O3 standards. 

One recent controlled human exposure study, reviewed in Chapter 6, included exposures to 

low, clearly environmentally-relevant, O3 concentrations for a group of healthy young adults. 

Adams (2006) reported that following a 6.6 h exposure to 0.08 ppm O3, group mean O3-induced 

FEV1 decreases (!6.1%, square-wave; !7.0%, triangular) and symptom responses in healthy 

adults were statistically significantly greater than after 0.04 and 0.06 ppm O3 exposures. During 

a 6.6 h exposure to 0.06 ppm O3, group mean FEV1 responses diverged from responses for 

filtered-air and 0.04 ppm O3 by 5.6 h, but did not reach statistical significance at p # 0.05 by the 

end of the exposure period (Adams, 2006). Exposures to 0.04 ppm O3 for 6.6 h produced group 

mean FEV1 responses quantitatively similar to those observed for filtered-air exposures (Adams, 

2002, 2006). These results suggest that acute exposure to O3 can have effects in healthy young 

adults with exposure levels below 0.08 ppm O3 (6.6 hr average) although no statistically 

significant effects were observed at the lowest exposure level of 0.04 ppm O3. It is important to 

note, however, that there is considerable interindividual variability in responses, as was 

illustrated in Figures 8-1B and 8-2, and some evidence of FEV1 decrements >10% for some 

subjects at 0.06 ppm. It is also important to note that this type of study generally includes 

healthy individuals and more subtle health measures, in contrast to population-based 

epidemiologic studies. 

8-42
The potential for the existence of threshold levels has been explored in several recent 

epidemiologic studies, as described in more detail in Section 7.6.5. A U.S. multicity panel study 

by Mortimer et al. (2002) examined associations of ambient O3 concentrations with changes 

in lung function and respiratory symptoms in asthmatic children. The mean 8-h avg O3 

(10 a.m.-6 p.m.) was 48 ppb, with less than 5% of study days exceeding 80 ppb. Analyses 

indicated that ambient O3 was significantly associated with decrements in PEF, both when using 

all data available and after restricting data to days when 8-h avg O3 concentrations were 

<80 ppb. A slightly larger effect estimate was observed using the restricted data set. Ozonerelated increased incidence of respiratory symptoms also persisted after eliminating data when 

8-h avg O3 exceeded 80 ppb. 

Examination of the shape of the concentration-response function in several time-series 

studies of O3-related ED visits and hospitalizations has provided some indications of an effect 

threshold. In a study of ED visits for asthma in St. John, Canada, the O3 effect observed in the 

>15 years age group was apparent only when 1-h max O3 data above the 95th percentile value of 

75 ppb were included (Stieb et al., 1996). In another study conducted in Toronto, Canada, the 

association between ambient O3 and all-age respiratory hospital admissions only became 

apparent above a daily 1-h max O3 level of approximately 30 ppb (Burnett et al., 1997b). In 

London, England, possible thresholds for O3 effects on respiratory hospitalizations were 

observed at 40 to 50 ppb for 8-h max O3 and 50 to 60 ppb for 1-h max O3 (Ponce de Leon et al., 

1996). However, other studies have reported a monotonic increase in the concentration-response 

function throughout the range of ambient O3 concentrations, suggesting a lack of any evident 

threshold for O3-related effects on respiratory hospitalizations and asthma ED visits (Burnett 

et al., 1997a; Jaffe et al., 2003; Petroeschevsky et al., 2001; Tenías et al., 1998). 

In the 95 U.S. communities study, Bell et al. (2004) reported that the risk estimate for 

excess mortality associated with short-term O3 exposure was slightly smaller but remained 

statistically significant when only using data from days with 24-h avg O3 concentrations below 

60 ppb. A more formal threshold analyses recently reported by Bell et al. (2006) for 98 U.S. 

communities, including the same 95 communities in Bell et al. (2004), indicated that if a 

population threshold existed for mortality, it would likely fall below a 24-h avg O3 concentration 

of 15 ppb. Other analyses by Kim et al. (2004) investigating the presence of a threshold in O3­

mortality effects in Seoul, Korea estimated threshold values of 28 ppb in 1-h max O3 when using 

8-43
all-year data and 45 ppb for summer-only data. None of the other pollutants examined, 

including PM10, SO2, NO2, and CO, had a nonlinear association with mortality. Results from the 

Kim et al. study suggested that if a threshold truly exists, the use of log-linear models may 

underestimate the O3 effect on mortality at levels above the threshold. 

Taken together, the available evidence from toxicologic, clinical and epidemiologic studies 

suggests that no clear conclusion can now be reached regarding possible threshold levels for O3­

induced effects. The controlled human exposure studies demonstrate notable variability in 

responsiveness among healthy subjects, even at low (#0.08 ppm) O3 exposure levels, making it 

infeasible to suggest any clear threshold for the heterogeneous general population from these 

results. As discussed in Section 7.6.5, there are limitations in epidemiologic studies that make 

discerning thresholds in populations difficult, including low data density in the lower 

concentration ranges, the possible influence of measurement error, and interindividual 

differences in susceptibility to O3-related effects in populations. Nevertheless, the limited 

clinical and epidemiologic evidence suggests that if a population threshold level does exist, it is 

likely near the lower limit of ambient O3 concentrations in the United States. 

8.6 BIOLOGICAL PLAUSIBILITY AND COHERENCE OF EVIDENCE 

FOR OZONE-RELATED HEALTH EFFECTS 

This section integrates findings from epidemiologic studies with toxicologic and 

mechanistic information obtained from controlled human exposure studies and animal 

toxicology studies for major health endpoints reported to be associated with either short- or longterm exposure to ambient O3. The section focuses on evidence related to two key considerations 

for drawing conclusions about causality – biological plausibility and coherence of the evidence. 

Evaluation of the biological plausibility of the O3-health effects associations observed in 

epidemiologic studies reflects consideration of both exposure-related factors and 

dosimetric/toxicologic evidence relevant to identification of potential biological mechanisms. 

Similarly, coherence of health effects associations reported in the epidemiologic literature 

reflects consideration of information pertaining to the nature of the various respiratory- and 

cardiovascular-related morbidity and mortality effects and physiological endpoints evaluated in 

human and animal toxicologic studies. The discussion in each subsection summarizes pertinent 

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key information and then discusses plausibility based on toxicological studies of effects 

attributed to O3 exposures in epidemiologic studies. 

For this assessment, the ensuing discussion on biological plausibility and coherence 

considers (a) the extent to which available epidemiological evidence logically ties together a 

range of relevant health endpoints (from cardiopulmonary physiological changes to morbidity to 

mortality) and (b) the extent to which available toxicological and biochemical evidence supports 

plausible causal relationships for observed epidemiological associations for specific types of 

health outcomes, i.e., how well do key epidemiologic findings compare with reasonably 

hypothesized or experimentally demonstrated biological mechanisms of action. 

8.6.1 Acute Ozone Exposure-Induced Health Effects 

Respiratory Health Effects 

As noted in Section 8.4.3, several new epidemiologic (field/panel) studies show positive 

associations between short-term exposure to ambient O3 and human respiratory effects. These 

health effects, as evaluated, include reductions in lung function, increased use of asthma 

medication, and increased hospitalization, especially among individuals with asthma or certain 

known cardiopulmonary diseases (see Chapter 7). The patterns of experimentally demonstrated 

physiological and biochemical alterations noted earlier (see Table 8-1 and Figures 8-3 and 8-4) 

support certain hypotheses regarding underlying pathological mechanisms in the development of 

respiratory effects reported in the epidemiologic studies. Some of these mechanisms (see Table 

8-1) include (a) decrements in lung function (capacities and volume), (b) bronchoconstriction, 

(c) increased airway responsiveness, (d) airway inflammation, (e) epithelial injury, (f) immune 

system activation, (g) host defense and (h) individual sensitivity factors such as age, genetic 

susceptibility and the extent of tolerance resulting from previous exposures. The time sequence, 

magnitude, and overlap of these complex events, both in terms of development and recovery (see 

Figures 8-3 and 8-4), indicate the difficulties associated with the interpretation of biological 

plausibility associated with the cardiopulmonary health effects. 

Controlled human exposure studies have clearly demonstrated the following three types of 

respiratory responses to acute O3 exposures: (1) irritative cough and substernal chest pain upon 

inspiration; (2) decrements in FVC and FEV1 due to decreased inspiratory capacity rather than 

airways obstruction and (3) neutrophilic inflammation of the respiratory tract. Increased 

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sensitivity (susceptibility) to these effects was observed even among a carefully selected 

homogeneous study population. The sources of this heterogeneity are not fully understood. 

As discussed in the earlier section, changes in baseline levels of various responses, the lag in the 

recovery phase and the role of residual defects in these mechanisms in hyperresponsive 

individuals suggest the potential for increased health effects in compromised individuals with 

preexisting cardiopulmonary diseases. Recent research has emphasized further characterization 

of the mechanisms and consequences of O3-induced pulmonary function and inflammatory 

responses. In addition, animal studies indicate morphological changes associated with acute O3 

exposures. 

Ozone-induced altered breathing patterns (rapid shallow breathing) observed in controlled 

human exposure studies and animals occur without significantly affecting minute ventilation, 

suggesting compensatory changes in breathing pattern. Such a shift in breathing pattern 

diminishes deep lung penetration of O3. Breathing pattern is modulated by changes in peripheral 

mechanisms, such as direct or indirect stimulation of lung receptors and bronchial C-fibers. The 

activity of these afferents is integrated with input from sensory pathways and thus determines the 

depth and frequency of breathing. Stimulation of bronchial C-fibers along with inhibition of 

inspiration through local axon reflexes can induce neurogenic inflammation via tachykinins and 

other proinflammatory neuropetides. Ozone-induced increases in the levels of the neuropeptide 

substance P observed in the BALF of human subjects suggests potential neurogenic involvement 

in O3-induced increased vascular permeability, plasma protein extravasation, mucus secretion, 

and bronchoconstriction (Solway and Leff, 1991). Similar neurogenic involvement due to 

vagally mediated stimulation of C-fibers seen in animal toxicology studies also supports O3­

induced airway hyperresponsiveness observed in humans. 

An extensive database of animal, human, and in vitro studies supports the conclusion 

that O3 interacts with airway epithelial cell membranes and lining fluid to form lipid ozonation 

products and reactive oxygen species (ROS). These reactive products initiate a cascade of 

events leading to oxidative stress, injury, inflammation, airway epithelial damage and increased 

alveolar permeability to vascular fluids. Inflammation is the outcome of host response to injury 

and usually resolves completely. Continued irritant challenge may evolve into a chronic 

inflammatory state with simultaneous alterations in lung structure and function, leading to 

diseases such as fibrosis and emphysema, although neither fibrosis nor emphysema have yet 

8-46
been observed with O3 exposure. Continued inflammation can also alter the lung’s ability to 

respond to infectious agents, allergens, and toxins. Acute inflammatory responses to O3 

exposure have been well documented experimentally in both humans and animals. As presented 

in Figure 8-4, the early inflammatory response to O3-induced lung injury is apparent in human 

subjects within 3 h postexposure. This initial neutrophilic inflammatory response phase is 

characterized by increases in PMNs in BALF along with increased levels of inflammatory 

mediators such as interleukins, prostaglandins and complement component C3a. In vitro studies 

using human and animal lung cell culture systems have further examined the involvement of 

various inflammatory mediators and in some instances their downstream signaling pathways. 

The late inflammatory phase in the lung is characterized by increased levels of monocytes 

and eosinophils, as well as respective mediators such as cytokines, leukotrines, proteinases, 

and ROS. 

Disruption of the lung’s blood barrier by O3 results in vascular permeability changes 

and plasma protein extravasation. Analysis of BALF plasma influx markers such as albumin, 

other proteins, immunoglobulins, and epithelial cell damage markers such as LDH indicate 

O3-induced lung epithelial injury. Ozone-induced lung injury and subsequent disruption of the 

airway epithelial barrier has been implicated in altered mucociliary clearance of particles 

observed in controlled human studies. Analogously, animal toxicology studies (see Chapter 5) 

have reported increased mortality to bacterial and viral infections subsequent to O3 exposure and 

also increased clearance of particles with low O3 exposure levels. 

Controlled O3 exposure studies of healthy humans have indicated a large degree of 

intersubject variability. The spirometric and symptomatic responses are highly reproducible 

within a given subject; but, within a group, pulmonary function can vary widely (see Figure 8-1) 

across different subjects. This interindividual variability is likely to be due to factors such as 

age, genetic background and antioxidant defenses, but as yet other factors remain to be 

characterized. Epidemiologic studies indicate a positive association between exposure to 

ambient O3 and declines in lung function in children and those with preexisting respiratory 

diseases such as asthma. Increased incidence of ED visits and hospitalizations due to respiratory 

causes that have been reported in various epidemiologic studies (discussed in Chapter 7), as also 

supported by animal toxicology data, suggest a causal association with O3. The epidemiologic 

associations become more apparent when the data are analyzed for the influence of seasonal 

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differences in ambient O3 levels, with stronger evidence for associations in the warm season. 

Together, the evidence from the animal and human studies suggests that acute O3 exposure is 

causally associated with respiratory system effects, including O3-induced pulmonary function 

decrements, respiratory symptoms, lung inflammation, and increased lung permeability, airway 

hyperresponsiveness, increased uptake of nonviable and viable particles, and consequent 

increased susceptibility to PM-related toxic effects and respiratory infections. 

Cardiovascular Health Effects 

Only a few experimental studies of animals and humans have evaluated possible 

mechanisms or physiological pathways by which acute O3 exposures may induce cardiovascular 

effects. Ozone induces lung injury, inflammation, and impaired mucociliary clearance, with a 

host of associated biochemical changes all leading to increased lung epithelial permeability. 

As discussed in Section 5.2.1, the generation of lipid ozonation products and reactive oxygen 

species in lung tissue can influence pulmonary hemodynamics and, ultimately, the 

cardiovascular system. 

Recent in vitro studies of O3 reactions with cholesterol in lung surfactant found consequent 

generation of highly reactive products such as oxysterols and $-epoxide in BALF isolated from 

rats exposed to 2.0 ppm O3 for 4 h (Pulfer and Murphy, 2004). Additionally, both 5$,6$-

epoxycholesterol and its most abundant metabolite, cholestan-6-oxo-3$,5"-diol, were shown to 

be cytotoxic to human lung epithelial (16-HBE) cells and to inhibit cholesterol synthesis. 

Studies (Pulfer et al., 2005) of mice exposed to 0.5, 1.0, 2.0, or 3.0 ppm O3 for 3 h also 

demonstrated that these oxysterols were produced in vivo. These results suggest that this may 

be an additional mechanism of O3 toxicity, including a pathway by which O3 may play a possible 

role in atherosclerosis and other cardiovascular effects. 

The presence of oxysterols in human atherosclerotic lesions implicates the oxidation of 

cholesterol in the pathogenesis of atherosclerosis, a well-known contributor to development of 

cardiovascular disease. Oxysterols may arise from different cholesterol oxidation mechanisms, 

(including free radical-mediated oxidations), and their unabated accumulation in macrophages 

and smooth muscle cells of arterial walls lead to formation of fatty streaks in advanced lesions. 

The presence of one of the O3-induced oxysterols, secosterol, in endogenously formed arterial 

plaques (Wentworth et al., 2003) suggests that the oxysterols produced in the lung either due to 

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direct O3 interaction with surfactant cholesterol or with oxidant radicals at the O3-induced 

inflammation site may have potential involvement in the development of cardiovascular and 

myocardial diseases. In addition, the recent in vitro observation (Sathishkumar et al. 2005) of 

increased apoptosis (programmed cell death) induced by secosterol in H9c2 cardiomyocytes 

(heart cells) supports possible involvement of such biologically active oxysterols in O3-induced 

cardiovascular effects observed in the epidemiologic studies. Also, the detection of oxysterols in 

the BALF of rats exposed to O3 suggests their potential to be used as biomarkers of O3 exposure. 

Demonstration of relationships between oxysterols of the type generated in lung surfactant with 

O3 exposure and cardiovascular disease outcomes in clinical settings or epidemiologic studies 

would add considerable value to the experimental observations thus for reported in the animal 

toxicology studies. 

Ozone-induced changes in heart rate, edema of heart tissue, and increased tissue and serum 

levels of atrial natriuretic factor (ANF) found with 8-h 0.5 ppm O3 exposure in animal 

toxicology studies (Vesely et al., 1994a,b,c) also raise the possibility of potential cardiovascular 

effects of acute O3 exposures. Such effects resulting from stimulation of airway irritant 

receptors, c-fiber activation, may result from either local or central nervous system involvement. 

Only one controlled human study (Gong et al., 1998) evaluated potential cardiovascular health 

effects of O3 exposure and reported O3 -induced changes in alveolar-arterial oxygen transfer in 

subjects with hypertension. However, several other cardiovascular parameters (e.g., heart rate) 

evaluated in this study did not show any O3-induced effects. 

Animal toxicology studies have found both transient and persistent ventilatory responses 

with or without progressive decrease in heart rate (Arito et al., 1997). Observations of 

O3-induced vasoconstriction in a controlled human exposure study by Brook et al. (2002) 

suggests another possible mechanism for O3-related exacerbations of preexisting cardiovascular 

disease. 

A few new field/panel studies of human adults have reported associations between ambient 

O3 concentrations and changes in electrophysiologic indicators of cardiac function, e.g., heart 

rate variability (HRV). Also, some population time-series studies have suggested positive 

associations between acute O3 exposure and cardiovascular hospitalizations in the warm season. 

These suggestive positive epidemiologic associations gain credibility and scientific support from 

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results of experimental animal toxicology and human clinical studies, as discussed above, which 

are indicative of plausible pathways by which O3 may exert cardiovascular effects. 

8.6.2 Chronic O3 Exposure-Induced Health Effects 

The effects of chronic O3 exposure in humans were earlier addressed primarily with crosssectional epidemiologic studies, as discussed in the 1996 O3 AQCD. Due to lack of precise 

information on exposure, the possibility of selection bias, and the difficulty of controlling for 

confounders, the earlier study findings were considered to be inconclusive. Since 1996, several 

new cross-sectional epidemiological studies have evaluated potential associations between 

chronic exposure to O3 and morbidity or mortality (see Chapter 7, Section 7.5). These studies 

suggest that seasonal exposure to O3 may be related to changes in lung function in children. 

However, little evidence is available to support a relationship between chronic O3 exposure and 

mortality or lung cancer incidence. There are no data available from controlled human chamber 

studies that evaluated chronic exposure regimens. 

The lack of adequate data from epidemiologic and clinical studies in human has helped to 

focus attention much more so on results from chronic O3 exposure studies in animals. Earlier 

chronic animal studies employed traditional exposure designs using chronic stable O3 exposures 

to one or another single O3 concentration over extended periods (weeks, months). More recent 

studies have attempted to incorporate design features that more closely mimic diurnal and 

seasonal patterns of O3 exposure and realistic exposure concentrations. Studies of monkeys that 

compared these two designs reported greater airway pathology with the latter design. Persistent 

and irreversible effects observed in chronic animal toxicology studies indicate the need for 

further complementary human data from epidemiologic studies. 

Animal toxicology data provide a clearer picture indicating that long-term O3 exposure 

may have lasting effects. Chronic exposure studies in animals have reported biochemical and 

morphological changes suggestive of irreversible long-term O3 impacts on the lung. Some of the 

studies in rats (0.5-1.0 ppm O3 for 6 h/day) for 20 months and monkeys (0.61 ppm) for one year 

noted increased deposition of collagen and thickening of the CAR. Differences in the degree of 

this type of lung damage have been observed between continuous exposure and seasonal patterns 

of variations in O3 exposure levels over time. A long-term study of infant rhesus monkeys 

exposed to simulated seasonal O3 (0.5 ppm 8 h/day for 5 days every 14 days for 11 episodes) 

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resulted in remodeling in the distal airways, abnormalities in tracheal basement membrane, 

accumulation of eosinophils in conducting airways and decrements in airway innervation. 

Earlier studies in rats following seasonal episodic profiles also showed small, but significant, 

decrements in lung function that were consistent with focal fibrinogenesis in the proximal 

alveolar region. On the other hand, chronic O3 exposures in a range of 0.5 to 1.0 ppm induce 

epithelial hyperplasia that disappears in a few days. The weight of evidence from the new 

experimental animal studies (using non-lifetime exposures) does not support ambient O3 as being 

a pulmonary carcinogen. 

Collectively, evidence from animal studies strongly suggests that chronic O3 exposure is 

capable of damaging the distal airways and proximal alveoli, resulting in lung tissue remodeling 

leading to apparent irreversible changes. Compromised pulmonary function and structural 

changes due to persistent inflammation may exacerbate the progression and development of 

chronic lung disease. These findings offer some insight into potential biological mechanisms for 

the suggested association between seasonal O3 exposure and reduced lung function development 

in children as observed in epidemiologic studies. 

8.6.3 Mortality-Related Health Endpoints 

An extensive analysis of population time-series studies that evaluated the air pollution 

related mortality risk estimates presented in Section 7-4 utilized data from single and multicity 

studies from around the world. Mortality risk estimates derived from single- and multicity 

studies in U.S. and Canada coupled with meta-analyses generally indicate associations between 

acute O3 exposure and elevated risk for all-cause mortality, even after adjustment for the 

influence of season and PM. Several single-city studies that specifically evaluated the 

relationship between O3 exposure and cardiopulmonary mortality also reported results suggestive 

of a positive association. 

The epidemiology results outlined above for mortality suggest a pattern of effects that may 

be biologically germane to interpretation of its causality, but our knowledge about potential 

underlying mechanisms remains relatively limited and suggests a need for further experimental 

support. The majority of the physiological and biochemical parameters evaluated both in human 

clinical and animal toxicology studies (Table 8-1; Figure 8-3) suggest a relatively transient 

nature for O3-induced biochemical perturbations. Most effects attenuate over time, depending on 

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the preexisting pathophysiology. However, one can hypothesize a generic pathway of O3­

induced lung damage, potentially involving oxidative lung damage with subsequent 

inflammation and/or decline in lung function leading to respiratory distress in some sensitive 

population groups (e.g., asthmatics) or, other plausible pathways noted below may lead to O3­

related contributions to cardiovascular effects that ultimately increase the risk of mortality. 

Recent analysis of data from the third National Health and Nutrition Examination 

Followup study indicated that about 20% of the adult population have reduced FEV1 values 

indicative of impaired lung function. The majority of these individuals have COPD, asthma or 

fibrotic lung disease (Mannino et al., 2003). These cardiopulmonary disease conditions are 

associated with persistent low-grade systemic inflammation. It has also been reported that 

patients with COPD are at increased risk for cardiovascular disease. Lung disease with 

underlying inflammation may also link to low-grade systemic inflammation associated with 

atherosclerosis. These effects in disease are independent of cigarette smoking (Sin et al., 2005). 

Lung function decrements in cardiopulmonary disease have also been associated with 

inflammatory markers such as C-reactive protein (CRP) in blood. In fact, at the population 

level, individuals with the lowest FEV1 have the highest levels of CRP, while those with highest 

FEV1 have the lowest values for CRP (Mannino et al., 2003; Sin and Man, 2003). The complex, 

physiological and biochemical perturbations that exist simultaneously (Figures 8-3 and 8-4) 

subsequent to acute exposure to O3 may tilt the biological homeostasis mechanisms leading to 

adverse health effects in people with compromised cardiopulmonary systems. 

Of much interest are several other types of newly available experimental data that support 

reasonable hypotheses that may help to explain findings of O3-related increases in cardiovascular 

mortality observed in some epidemiological studies. These include the direct effects of O3 in 

terms of its increasing platelet aggregating factor (PAF) in lung tissue that can then enter the 

general circulation and possibly contribute to increased risk of blood clot formation (as 

illustrated in Figure 5-3) and consequent increased risk of myocardial infarction (heart attack), 

cerebrovascular events (stroke), and/or associated cardiovascular-related mortality. Other O3­

induced effects may also contribute to increased risk of cardiovascular impacts. Ozone reactions 

with cholesterol in lung surfactant to produce epoxides and/or associated diol metabolites that 

are cytotoxic to lung and heart muscle cells and that contribute to atherosclerotic plaque 

formation in arterial walls represent another such pathway. Also, to the extent that O3-induced 

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increases in lung permeability allow more ready entry of inhaled particulate matter (PM) into the 

blood stream, then O3 exposure would enhance the risk for PM-related cardiovascular effects 

(see U.S. Environmental Protection Agency, 2004). An associated factor to consider is O3’s 

ability to contribute to ultrafine PM formation in ambient air and indoor environments (as 

discussed in Sections 8.2.4 and 8.3.2). Thus, O3 can both contribute to increased presence of 

fine particles and human exposure to them, as well as enhance uptake of inhaled fine PM and 

thereby presumably contribute to exacerbation of PM-induced cardiovascular effects in addition 

to those more directly induced by O3 per se. 

8.6.4 Health Effects of Ozone-Containing Pollutant Mixtures 

The above-noted potential for O3-related enhancements of PM formation, particle uptake, 

and exacerbation of PM-induced cardiovascular effects underscores the importance of 

considering contributions of O3 interactions with other often co-occurring air pollutants to health 

effects due to O3-containing pollutant mixes. Chapters 4, 5, and 6 provided a discussion of 

experimental studies that evaluate interactions of O3 with other co-occurring pollutants. Some 

examples of important pollutant mixture effects noted there are highlighted below. 

First, Chapter 4 noted some important interactive effects of coexposures to O3 and NO2 

and SO2, two other common gaseous copollutants found in ambient air mixes. That is, a study 

by Rigas et al. (1997) showed that continuous exposure of healthy human adults to SO2 or 

to NO2 increased inhaled bolus O3 absorption, while continuous exposure to O3 alone decreased 

bolus absorption of O3. This suggests enhancement of O3 uptake by NO2 or SO2 coexposure in 

ambient air mixes. Also, as noted in Chapter 6, another study by Jenkins et al. (1999) showed 

that asthmatics exhibited enhanced airway responsiveness to house dust mite following 

exposures to O3, NO2, and the combination of the two gases. Spirometric response, however, 

was impaired only by O3 and O3+NO2 at higher concentrations. On the other hand, animal 

toxicology studies discussed in Chapter 5 that evaluated exposures to O3 in mixture with NO2, 

formaldehyde, and PM demonstrated additive, synergistic or antagonistic effects, depending on 

the exposure regimen and the specific health endpoints evaluated. 

The results of the Jenkins et al. (1999) study also help to illustrate a more general 

phenomena of enhancement by O3 exposure of various respiratory responses of sensitive 

individuals to allergens. Chapter 6 noted, for example, studies (a) by Peden et al. (1995) 

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showing O3-induced increased response to nasal allergen challenge among allergic asthmatic 

subjects, and (b) by Michelson et al. (1999) showing promotion by 0.4 ppm O3 exposure of 

inflammatory cell influx in response to nasal allergen challenge in asymptomatic dust-mite 

sensitive asthmatics. In addition, Jörres et al. (1996) demonstrated enhancement by 0.25 ppm O3 

exposure of airway responsiveness in mildly allergic asthmatics that was increased in response to 

an individual’s historical allergen (grass and birch pollen, house dust mite, animal dander). 

These results were further extended by Holz et al. (2002) who showed that repeated daily 

exposure to 0.125 ppm O3 for 4 days exacerbated lung function decrements (e.g., decreased 

FEV1) in response to bronchial allergen challenges among subjects with preexisting allergic 

airway disease, with or without asthma. This suggests that O3 exposure can place allergic 

nonasthmatic persons, as well as asthmatics, at increased risk for allergic respiratory effects. 

Consistent with and supporting the above findings are animal toxicology studies reviewed in 

detail by Harkema and Wagner (2005), which indicate that (a) O3-induced epithelial and 

inflammatory responses in laboratory rodents are markedly enhanced by coexposure to inhaled 

biogenic substances (e.g., bacterial endotoxin or ovalbumin, an experimental aeroallergen) and 

(b) adverse airway effects of biogenic substances can be exacerbated by coexposure to O3. 

Also of much note is a newly emerging literature which indicates that O3 can modify the 

biological potency of certain types of ambient PM, as shown by experimental tests. For 

example, as described in Chapter 5 (Section 5.4.2) reaction of diesel PM with 0.1 ppm O3 for 

48 h increased the potency (compared to non-exposed or air-exposed diesel PM) to induce 

neutrophil inflex, total protein, and LDH in lung lavage fluid in response to intratracheal PM 

instillation in rats (Madden et al., 2000). However, the potency of carbon black particles was not 

enhanced by exposure to O3, suggesting that O3 reaction with organic components of the diesel 

PM were responsible for the observed increased diesel PM effects. 

Potential interaction of O3 with fine PM in aged rats was examined by Kleinman et al. 

(2000). In this study the effects of fine PM containing two common toxic constituents, 

ammonium bisulfate (ABS, 0.3 μm 70 μg/m!3

) and elemental carbon (C, 0.3 μm 50 μg/m!3) and 

a mixture (ABS + C) with 0.2 ppm O3 was evaluated on aged rat lung structure and macrophage 

function. Exposures of O3, elemental carbon or ABS alone did not cause significant lung injury, 

lung tissue collagen content or respiratory burst activity. On the other hand, mixtures (ABS + 

C + O3) caused significant lung injury as assessed by increased cell proliferation response in 

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lung epithelial and interstitial cells, loss of lung tissue collagen and increase in respiratory burst 

and phagocytic activity. 

The majority of toxicological studies discussed earlier in this chapter evaluated effects of 

individual pollutants or simple mixtures of the constituents of urban smog mixtures, and these 

toxicology studies do not fully explain epidemiologic findings that have increasingly shown 

ambient O3, other gaseous pollutants, and/or PM to be associated with various health effects at 

relatively low concentrations. In a recent report, Sexton et al (2004) utilized “smog chambers”, 

i.e., environmental irradiation chambers to generate synthetic photochemical oxidants mixtures 

similar to urban smog, and studied the toxicity of such mixtures on the inflammatory response of 

A549 cells in an in vitro exposure system. In this preliminary study, the authors found the 

simulated urban photochemical oxidant mixture generated with the addition of O3 to have 

enhanced toxicity (as assessed by the expression of IL-8 mRNA). Additional toxicology studies 

using similar realistic air pollution smog mixtures in the future may provide more relevant 

biological understanding for the potential interactions that occur in the ambient air among 

various pollutants. 

All of the above types of interactive effects of O3 with other co-occurring gaseous and 

nongaseous viable and nonviable PM components of ambient air mixes argue for not only being 

concerned about direct effects of O3 acting alone but also the need for viewing O3 as a surrogate 

indicator for air pollution mixes which may enhance risk of adverse effects due to O3 acting in 

combination with other pollutants. Viewed from this perspective, epidemiologic findings of 

morbidity and mortality associations with ambient O3 concentrations extending to concentrations 

below 0.08 ppm become more understandable and plausible. 

8.7 SUSCEPTIBLE AND VULNERABLE POPULATIONS, AND 

POTENTIAL PUBLIC HEALTH IMPACTS 

Many factors such as age, gender, disease, nutritional status, smoking, and genetic 

variability may contribute to the differential effects of environmental pollutants, including O3. 

Genetic factors, such as single nucleotide polymorphisms (SNPs) and developmental defects, 

can contribute to innate susceptibility, while acquired susceptibility may develop due to personal 

habits (smoking, diet, exercise) and other risk factors such as age, gender, pregnancy, and 

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copollutant exposures. In the 1996 O3 AQCD, children, outdoor workers, and people with 

preexisting respiratory disease were identified as likely being more susceptible or vulnerable to 

effects of ambient O3 exposure. However, the available toxicological and human data had not 

shown that males and females respond differently to O3; nor were available data adequate to 

suggest differences in responsiveness to O3 based on ethnic or racial background. Overall, then, 

the available information from animal toxicology and epidemiologic studies provided only 

relatively limited evidence by which to confidently identify likely susceptible groups and/or to 

associate specific factors as contributing to increased risk of O3-related adverse health effects. 

Advances in available research results since 1996 have improved our ability to delineate likely 

susceptible or vulnerable populations at increased risk for O3-induced health effects and to 

delineate factors contributing to such risk. 

8.7.1 Preexisting Disease as a Potential Risk Factor 

People with preexisting pulmonary disease are likely to be among those at increased risk 

from O3 exposure. Altered physiological, morphological and biochemical states typical of 

respiratory diseases like asthma, COPD and chronic bronchitis may render people sensitive to 

additional oxidative burden induced by O3 exposure. Based on studies assessed in the 1996 O3 

AQCD (U.S. Environmental Protection Agency, 1996a), asthmatics appear to be at least as, or 

more, sensitive to acute O3 exposure as healthy nonasthmatic subjects. The new results reviewed 

in Chapters 6 and 7 of this document from controlled exposure and epidemiologic studies also 

indicate that asthmatics are a potentially sensitive subpopulation for O3 health effects. 

A multicity study by Mortimer et al. (2002) support earlier observations that asthmatic 

children are particularly susceptible to ambient O3. This association, based on decrements in 

lung function and exacerbation of pulmonary disease symptoms, suggests that O3 exposures may 

result in increased use of medication in asthmatic children. A number of time-series 

epidemiologic studies have reported increased risk in study subsets of individuals with 

preexisting lung diseases, especially asthma, as potentially susceptible individuals. The 

epidemiologic studies of acute exposure to O3 discussed in Section 8.4.2 indicate increased risk 

for exacerbation of respiratory disease symptoms during the warm season. 

Several clinical studies reviewed in the 1996 O3 AQCD on atopic and asthmatic subjects 

had suggested but not clearly demonstrated enhanced responsiveness to acute O3 exposure 

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compared to healthy subjects. The majority of the newer studies reviewed in Chapter 6 indicate 

that asthmatics are as sensitive as, if not more sensitive than, normal subjects in manifesting O3­

induced pulmonary function decrements. 

Ozone-induced increases in neutrophils, protein, and IL-8 were found to be significantly 

higher in the BALF from asthmatics compared to healthy subjects, suggesting mechanisms for 

the increased sensitivity of asthmatics. Similarly, subjects with allergic asthma exhibited 

increased airway responsiveness to inhaled allergens upon acute O3 exposure. Consistent with 

these changes, it is suggested that asthmatics will be more sensitive to small airway effects of 

ambient O3. Asthmatics present a differential response profile for cellular, molecular, and 

biochemical parameters (Figure 8-1) that are altered in response to acute O3 exposure. Increases 

in O3-induced nonspecific airway responsiveness incidence and duration could have important 

clinical implications for asthmatics. 

Bronchial constriction following provocation with allergens presents a two-phase response. 

The early response is mediated by release of histamine and leukotrienes that leads to contraction 

of smooth muscle cells in the bronchi, narrowing the lumen and decreasing the airflow. 

In asthmatics, these mediators also cause accumulation of eosinophils, followed by production 

of mucus and a late-phase bronchial constriction and reduced airflow. Holz et al. (2002) 

reported an early phase response in subjects with rhinitis after a consecutive 4-day exposure to 

0.125 ppm O3 that resulted in a clinically relevant (>20%) decrease in FEV1. Allergen challenge 

in mild asthmatics 24 h postexposure to 0.27 ppm O3 for 2 h resulted in significantly increased 

eosinophil counts in BALF compared to healthy subjects (Vagaggini et al., 2002). Epithelial 

cells from mucosal biopsies of allergic asthmatics indicated significant increases in the 

expression of IL-5, IL-8 and GM-CSF, suggesting increased neutrophilic inflammation 

compared to healthy subjects (Bosson et al., 2003). 

Several human exposure studies have shown differences between asthmatics and healthy 

human subjects with regard to PMN influx in BALF . In vitro studies (Schierhorn et al., 1999) 

of nasal mucosal biopsies from atopic and nonatopic subjects exposed to 0.1 ppm O3 found 

significant differences in release of IL-4, IL-6, IL-8, and TNF-". Another study by Schierhorn 

et al. (2002) found significant differences in the O3-induced release of the neuropeptides 

neurokinin A and substance P for allergic patients in comparison to nonallergic controls, 

suggesting increased activation of sensory nerves by O3 in the allergic tissues. Another study by 

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Bayram et al. (2002) using in vitro culture of bronchial epithelial cells recovered from atopic and 

nonatopic asthmatics also found significant increases in epithelial permeability in response to O3 

exposure. In addition, some controlled human O3 exposure studies in asthmatics (Hiltermann 

et al., 1999; Scannell et al., 1996) reported increased secretion of IL-8, suggesting increased 

neutrophilic inflammation. Two studies (Jörres et al., 1996; Holz et al., 2002) observed 

increased airway responsiveness to repeated daily O3 exposure to bronchial allergen challenge in 

subjects with preexisting allergic airway disease. 

Newly available reports from controlled human exposure studies (see Chapter 6) utilized 

subjects with preexisting cardiopulmonary diseases such as COPD, asthma, allergic rhinitis, and 

hypertension. The data generated from these studies that evaluated pulmonary function changes 

in spirometry did not find clear differences between filtered air and O3 exposure in COPD and 

asthmatic subjects. However, the new data on airway responsiveness, inflammation, and various 

molecular markers of inflammation and bronchoconstriction indicate that people with atopic 

asthma and allergic rhinitis comprise susceptible groups for O3-induced adverse health effects. 

Collectively, these observations suggest that O3 exposure exacerbates pre-existing allergic 

asthma. People with allergic asthma likely represent a sizable segment of the population 

reported to have increased symptoms of respiratory illness exacerbations, ED visits, and hospital 

admissions in epidemiologic studies. 

Although controlled human exposure studies have not found evidence of larger spirometric 

changes in people with COPD relative to healthy subjects, this may be due to the fact that most 

people with COPD are older adults who would not be expected to have such changes based on 

their age. However, new epidemiological evidence indicates that people with COPD may be 

more likely to experience other effects, including emergency room visits, hospital admissions, or 

premature mortality. A study by Tenías et al. (2002) observed a significant positive association 

between O3 and for COPD. Results from an analysis of five European cities indicated strong and 

consistent O3 effects on unscheduled respiratory hospital admissions, including COPD 

(Anderson et al., 1997). Also, an analysis of a 9-year data set for the whole population of the 

Netherlands provided risk estimates for more specific causes of mortality, including COPD 

(Hoek et al., 2000, 2001; reanalysis Hoek, 2003); a positive, but nonsignificant excess risk of 

COPD-related mortality was found to be associated with short-term O3 concentrations. 

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8.7.2 Age-Related Variations in Susceptibility/Vulnerability 

Growth and development of the human respiratory system is not complete until 18 to 

20 years of age; 80% of alveoli are formed postnatally, and changes in the lung continue through 

adolescence (Dietert et al., 2000). Because the developing lung is highly sensitive to damage 

from exposure to environmental toxicants, children and infants are likely to be among the most 

susceptible to detrimental effects from O3 exposure. In addition, children experience enhanced 

O3 vulnerability because they generally spend more time outdoors and are highly active, 

especially during the warm season when ambient O3 concentrations are expected to be high and 

results in increased O3 dose delivered to their lungs (Plunkett et al., 1992; Wiley et al., 1991a,b). 

Controlled human exposure studies assessed in the 1996 O3 AQCD indicated that children 

(generally aged 8 to 11 years) and adolescents (generally aged 12 to 18 years) exhibited, on 

average, greater spirometric responses to O3 than middle-aged or older adults when exposed to 

comparable O3 doses. However, less evidence of respiratory symptoms has been observed 

among healthy children in clinical studies of O3 exposure. Such diminished symptomatic 

responses in children may put them at an increased risk for continued O3 exposure. As reported 

in the 1996 O3 AQCD, several summer camp panel studies collectively provided strong evidence 

for an association between acute O3 exposure and lung function declines in mostly healthy, 

nonasthmatic children (aged 7 to 17 years) who spent long hours outdoors. Additional 

epidemiologic field studies observed that respiratory symptoms (or exacerbation of asthma) and 

decrements in PEF were associated with increased ambient O3 concentrations, with greater 

responses found in asthmatic than in nonasthmatic children. No new human exposure studies 

investigating O3 responses in children have been published since the last O3 AQCD; however, 

the epidemiologic studies published during the last decade (see Section 7.2.3.1 for details) 

continue to indicate that children are at increased risk for to O3-related respiratory health effects. 

Many recent field studies published in the past decade have focused on the effects of 

acute O3 exposure on the respiratory health of children. In general, children experienced 

O3-related decrements in pulmonary function parameters, including PEF, FEV1, and FVC (e.g., 

Mortimer et al., 2002; Thurston et al., 1997). Declines in lung function were observed in both 

healthy and asthmatic children following acute exposure to O3, but respiratory symptoms were 

largely found only in asthmatic children. The O3-related changes in lung function in children 

were often small, <5% change per 40 ppb increase in mean 1-h max O3 or equivalent. However, 

8-59
as stated earlier, the group mean responses reported may mask larger lung function changes in 

sensitive individuals. Höppe et al. (2003) examined several susceptible populations, including 

children (aged 6 to 8 years) and juvenile asthmatics (aged 12 to 23 years), for changes in 

pulmonary function attributable to O3 exposure in Munich, Germany. Using group mean values, 

consistent O3 effects were not observed in both groups; but, when data was analyzed on an 

individual basis, a potential pattern of increased O3 sensitivity was observed. About 20% of the 

children and asthmatics were found to be particularly susceptible to O3 health effects (i.e., 

experienced greater than 10% change in FEV1). 

Several epidemiologic field studies have examined effects of long-term O3 exposure on 

lung function in children. The results collectively indicate that seasonal O3 exposure is 

associated with smaller increases in lung function in children (mean age 7 to 8 years). These 

results are supported by a long-term study of infant rhesus monkeys exposed to simulated 

seasonal O3 patterns (0.5 ppm 8h/day for 5 days, every 14 days for 11 episodes) which found 

remodeling in the distal conducting airways as a result of damage and repair processes occurring 

with repeated O3 exposure (Evans et al., 2003; Schelegle et al., 2003). However, epidemiologic 

evidence of an effect of longer-term (i.e., multiyear) O3 exposure on lung function development 

in children has generally not yet been found, as most definitively indicated by the southern 

California Children’s Health Study that examined children from 10 to 18 years of age 

(Gauderman et al., 2004). 

In time-series studies that examined associations between ambient O3 concentrations and 

ED visits, hospitalizations, and mortality in children, consistent results were not observed. This 

is likely due to the small number of daily deaths and hospitalizations among children and, also, 

the fact that the large number of competing causes of mortality and hospitalizations in children 

(which are together more important than air pollution) make it difficult to detect any excess risk 

attributable to ambient O3 exposure. 

The elderly are also often classified as being particularly susceptible to air pollution. The 

basis for increased O3 sensitivity among older adults is not known, but one hypothesis is that it 

may be related to changes in the respiratory tract lining fluid antioxidant defense network (Kelly 

et al., 2003). Increased susceptibility of older adults to O3 health effects is most clearly indicated 

in the newer mortality studies. In the meta-analysis by Bell et al. (2005), a comparison of O3­

mortality risk estimates by age indicates that, in general, the older adults population (>65 years 

8-60
of age) is more susceptible to O3 effects, with older adults experiencing a more than 50% greater 

excess risk in mortality compared to the all-age group. In the large U.S. 95 communities study 

(Bell et al., 2004), the O3 effect on all-cause mortality was slightly larger for older adults 

compared to younger individuals (#65 years of age). Bell et al. (2004) further noted that though 

the risk estimates for older adults were only slightly higher, the absolute effect of O3 is 

substantially greater due to the higher underlying mortality rates in the older adult population 

(e.g., approximately 50% of deaths were in those $75 years), which leads to a larger number of 

extra deaths for older adults compared to the general population. 

Lung function responses to O3 exposure, however, have been found to be diminished in the 

older adult population. The study by Höppe et al. (2003) observed that ambient O3 exposure was 

not associated with lung function declines in seniors (69 to 95 years of age). While a greater 

than 10% change in FEV1 was observed in 20% of the children and juvenile asthmatics, only 5% 

of the older adult population experienced such a change. The results by Höppe et al. are 

consistent with the diminishing lung function responses to O3 exposure with increasing age 

observed in clinical studies. 

In summary, new epidemiologic research presented in Chapter 7 of this document 

continues to indicate that children, asthmatics in particular, constitute a sensitive group in 

epidemiologic studies of oxidant air pollution. Healthy children with prolonged exposure 

periods, combined with elevated levels of exertion or exercise also appear to be more vulnerable 

to O3-related respiratory health effects. Older adults (>65 years of age) also have been shown to 

be susceptible to O3 health effects, particularly O3-related mortality. One epidemiologic study 

also found that the elderly experience diminished lung function responses to O3 exposure, which 

is supported by evidence from clinical studies. 

8.7.3 Vulnerability of Outdoor Workers and Others Who Participate in 

Outdoor Activities 

The 1996 O3 AQCD indicated that one population group that was shown to have increased 

responsiveness to ambient O3 exposure consisted of exercising healthy and asthmatic 

individuals, including children, adolescents, and adults. The effects of O3 on the respiratory 

health of outdoor workers and others who participate in outdoor activities have been investigated 

in several recent epidemiologic studies. These individuals may experience increased 

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vulnerability for O3 health effects, because they are typically exposed to high doses of O3 as they 

spend long hours outdoors often at elevated exertion levels. 

In a group of berry pickers in Fraser Valley, Canada, large decrements in lung function 

(~5% decrease in FEV1 per 40 ppb increase in 1-h max O3) were associated with acute exposure 

to O3 (Brauer et al., 1996). The mean ambient 1-h max O3 was 40.3 ppb (SD 15.2) over the 

study period of June to August 1993. The berry pickers worked outdoors for an average of 11 h 

at elevated heart rates (on average, 36% higher than resting levels). These results indicate that 

extended exposures to O3 at elevated exertion levels can produce marked effects on lung 

function among outdoor workers. 

Höppe et al. (1995) examined forestry workers for O3-related changes in pulmonary 

function in Munich, Germany. Ventilation rates, estimated from their average activity levels, 

were elevated. When comparisons were made between high O3 days (mean 1⁄2-h max O3 of 64 

ppb) and low O3 days (mean 1⁄2-h max O3 of 32 ppb), 59% of the forestry workers experienced a 

notable decrement in lung function (i.e., at least a 20% increase in specific airway resistance or 

at least a 10% decrease in FEV1, FVC, or PEF) on high O3 days. None experienced improved 

lung function. This study also examined athletes following a 2-h outdoor training period in the 

afternoon yielding a ventilation rate double the estimate for the forestry workers. Though a 

significant association between ambient O3 levels and decrements in FEV1 was observed overall, 

a smaller percentage of the athletes (14%) experienced a notable decrement in lung function on 

high O3 days compared to the forestry workers; and 19% of the athletes actually showed an 

improvement. 

A large field study by Korrick et al. (1998) examined the effects of multihour O3 exposures 

(on average, 8 h) on adults hiking outdoors in Mount Washington, NH. The mean of the hourly 

O3 concentrations during the hike was 40 ppb (range 21-74). After the hike, all subjects 

combined experienced a relatively small mean decline in FEV1 (1.5% decrease per 30 ppb 

increase in mean hourly O3 concentrations) during the hike. Ozone-related changes in lung 

function parameters were estimated. Stratifying the data by hiking duration indicated that 

individuals who hiked 8 to 12 h experienced a >2-fold decline in FEV1 versus those only hiking 

2 to 8 h. 

Results from the above field studies are consistent with those from earlier summer camp 

studies (Avol et al., 1990; Higgins et al., 1990; Raizenne et al., 1987, 1989; Spektor et al., 1988, 

8-62
1991), which also observed strong associations between acute O3 exposure and decrements in 

lung function among children who spent long hours outdoors. In a recent analysis by the 

southern California Children’s Health Study, a total of 3,535 initially nonasthmatic children 

(ages 9 to 16 years at enrollment) were followed for up to 5 years to identify new-onset asthma 

cases associated with higher long-term ambient O3 concentrations (McConnell et al., 2002). 

Communities were stratified by pollution levels, with six high-O3 communities (mean 1-h max 

O3 of 75.4 ppb [SD 6.8] over four years) and six low-O3 communities (mean 50.1 ppb 

[SD 11.0]). In the combined analysis using all children, asthma risk was not found to be higher 

for residents of the six high-O3 communities versus those from the six low-O3 communities. 

However, within the high-O3 communities, asthma risk was more than 3 times greater for 

children who played three or more sports versus those who played no sports, an association not 

observed in the low-O3 communities. Therefore, among children repeatedly exposed to higher 

O3 levels, increased exertion outdoors (and resulting increased O3 dose) was associated with 

excess asthma risk. 

These field studies with subjects at elevated exertion levels support the extensive evidence 

derived from controlled human exposure studies. The majority of human chamber studies have 

examined the effects of O3 exposure in subjects performing continuous or intermittent exercise 

for variable periods of time (see Chapter 6). Significant O3-induced respiratory responses have 

been observed in clinical studies of exercising individuals. The epidemiologic studies discussed 

above also indicate that prolonged exposure periods, combined with elevated levels of exertion 

or exercise, may magnify O3 effects on lung function. Thus, outdoor workers and others who 

participate in higher exertion activities outdoors during the time of day when high peak O3 

concentrations occur appear to be particularly vulnerable to O3 effects on respiratory health. 

8.7.4 Genetic Factors Affecting O3 Susceptibility 

In the 1996 O3 AQCD, the potential influence of genetic factors on responses to O3 

exposure could not be thoroughly evaluated due to the very limited data then available. The 

human and toxicologic data were inadequate to suggest differences in O3 responsiveness based 

on gender, ethnic, or racial background. New animal toxicology studies using various strains of 

mice and rats that have identified O3-sensitive and resistant strains illustrate the importance of 

genetic background in determining O3 susceptibility. Using subacute low exposure regimen 

8-63
(0.3 ppm O3, 48 h) studies on inbred strains that have been designated as inflammation prone or 

resistant, Kleeberger et al., (1997) identified the pro-inflammatory cytokine gene, Tnf-", as a 

susceptibility gene. Further characterization of this model indicated a role for TNF receptors in 

O3-induced pulmonary epithelial injury and inflammation (Cho et al., 2001). Studies of five 

inbred strains of mouse with differing response to O3 exposure (acute high dose or low dose 

continuous exposure for 3 days), found a protective role for clara cell secretory protein against 

O3-induced oxidative damage (Broeckaert et al., 2003; Wattiez et al., 2003). The role for these 

genes and/or their orthologs in human susceptibility to O3 exposure has yet to be examined. 

Biochemical and molecular parameters evaluated in these toxicology experiments were 

used to identify specific loci on chromosomes and, in some cases, to relate the differential 

expression of specific genes to biochemical and physiological differences seen among these 

species. Using O3-sensitive and O3-resistant species, it has been possible to identify the 

involvement of AHR and inflammation processes in O3 susceptibility. However, most of these 

studies were carried out using relatively high doses of O3, making the relevance of these studies 

questionable for human health effects assessment. The molecular parameters identified in 

animal and human experimental studies may serve as useful biomarkers with the availability 

of suitable technologies and may, ultimately, be integrated with epidemiologic studies. The 

interindividual differences in O3 responsiveness observed across a spectrum of symptoms and 

lung function responses do not yet allow identification of important underlying factors, except a 

significant role for age. 

Apart from age at the time of exposure, controlled human exposure studies have also 

indicated a high degree of interindividual variability in some pulmonary physiological 

parameters. Genetic factors likely contribute to the substantial variability observed between 

individuals. Several recent human clinical and epidemiologic studies (Bergamaschi et al., 2001; 

Corradi et al., 2002; David et al., 2003; Romieu et al., 2004; Yang et al., 2005) have reported 

that genetic polymorphisms for antioxidant enzymes and inflammatory genes may modulate 

pulmonary function and inflammatory responses to O3 exposure. Glutathione S-transferases 

(GSTs) play a major role in protecting cells against damage from reactive oxygen species by 

conjugating them with glutathione so that they can be rapidly eliminated. A common 

homozygous deletion polymorphism of the GSTM1 gene (GSTM1 null genotype) abolishes 

enzyme activity and may increase susceptibility to O3-induced oxidative stress. Asthmatic 

8-64
children with a genetic deficiency of GSTM1 were found to be more responsive to ambient O3 

exposure, as assessed by decrements in FEF25-75, in a Mexico City study by Romieu et al. (2004). 

Antioxidant supplementation of vitamins C and E attenuated post-O3 lung function response in 

these children. More specific genotyping has shown that O3 responsiveness of asthmatic 

children may be related to the presence of variant Ser allele for a detoxifying enzyme (NQO1) 

induced in response to oxidative stress (David et al., 2003). The presence of at least one NQO1 

Ser allele in asthmatic children with GSTM1 null genotype was found to have a protective effect 

against O3 exposure. Also, polymorphism in Tnf-" has been implicated in the risk of O3-induced 

lung function changes in healthy, mild asthmatics and individuals with rhinitis (Yang et al., 

2005), which supports toxicologic evidence referred to earlier in this section. 

The above observations suggest a potential role for these genetic markers in the innate 

susceptibility to O3; however, the validity of these markers and their relevance in the context of 

prediction to population studies need additional validation. The lack of correlation between lung 

function and airway inflammatory responses to O3 in healthy subjects, combined with the 

evidence of separate chromosomal loci for O3-induced AHR and airway inflammation in inbred 

mice, suggest that these two responses are probably independently regulated. 

8.7.5 Potential Public Health Impacts 

Exposure to ambient O3 is associated with a variety of health outcomes, including 

increased incidence of cough, reduction in lung function, increased inflammation, and increased 

hospital admissions and mortality. In protecting public health, a distinction must be made 

between health effects that are considered “adverse” and those that are not. What constitutes an 

adverse health effect varies for different population groups, with some changes in healthy 

individuals not being viewed as adverse but those of similar type and magnitude in other 

susceptible individuals with preexisting disease being seen as adverse. 

8.7.5.1 Concepts Related to Defining of Adverse Health Effects 

The American Thoracic Society (ATS) published an official statement on “What 

Constitutes an Adverse Health Effect of Air Pollution?” (ATS, 2000). This statement updated 

guidance for defining adverse respiratory health effects published fifteen years earlier (ATS, 

8-65
1985), in order to take into account new investigative approaches used to identify the effects of 

air pollution and to reflect concern for impacts of air pollution on specific susceptible groups. 

In the 2000 update, there is an increased focus on quality of life measures as indicators of 

adversity and, also, a more specific consideration of population risk. Exposure to air pollution 

that increases the risk of an adverse effect to the entire population is viewed as adverse, even 

though it may not increase the risk of any identifiable individual to an unacceptable level. 

For example, a population of asthmatics could have a distribution of lung function such that no 

identifiable single individual has a level associated with significant impairment; and exposure to 

air pollution could shift the distribution to lower levels that still do not bring any identifiable 

individual to a level that is associated with clinically relevant effects. However, this would be 

considered to be adverse because individuals within the population would have diminished 

reserve function and, therefore, would be at increased risk if affected by another agent. 

Reflecting new investigative approaches, the ATS statement also describes the potential 

usefulness of research into the genetic basis for disease, including responses to environmental 

agents that provide insights into the mechanistic basis for susceptibility and provide markers of 

risk status. Likewise, biomarkers that are indicators of exposure, effect or susceptibility may 

someday be useful in defining the point at which one or an array of responses should be 

considered an adverse effect. 

The 1996 O3 AQCD provided information useful in helping to define adverse health effects 

associated with ambient O3 exposure by describing the gradation of severity and adversity of 

respiratory-related O3 effects, and those definitions are reproduced and presented here as 

Tables 8-2 and 8-3. The severity of effects described in those tables and the approaches taken to 

define their relative adversity still appear to be valid and reasonable even in the context of the 

new ATS statement (ATS, 2000). 

As assessed in detail in earlier chapters of this document and briefly recapitulated in 

preceding sections of this chapter, exposures to a range of O3 concentrations have been reported 

to be associated with increasing severity of several categories of health effects. 

Respiratory effects associated with short-term O3 exposures have been by far the most 

extensively studied and most clearly shown to be causally related to O3 exposure. Such effects 

include the induction of pulmonary function decrements and respiratory symptoms demonstrated 

in response to controlled acute (1- to 8-h) O3 exposures of human subjects and also observed 

8-66
Table 8-2. Gradation of Individual Responses to Short-Term Ozone Exposure in 

Healthy Persons a 

Functional Response None Small Moderate Large 

FEV1 Within normal Decrements of Decrements of Decrements of 

range (±3%) 3 to #10% >10 but <20% $20% 

Nonspecific Within normal Increases of Increases of #300% Increases of >300% 

bronchial responsiveness a range <100% 

Duration of response None <4 hours >4 hours but >24 hours 

#24 hours 

Symptomatic Response 

Cough 

Chest pain 

Duration of response 

Impact of Responses 

Normal 

Infrequent 

cough 

None 

None 

Normal 

Mild 

Cough with deep 

breath 

Discomfort just 

noticeable on 

exercise or 

deep breath 

<4 hours 

Normal 

Moderate 

Frequent 

spontaneous cough 

Marked discomfort 

on exercise or deep 

breath 

>4 hours but 

#24 hours 

Mild 

Severe 

Persistent 

uncontrollable 

cough 

Severe discomfort 

on exercise or 

deep breath 

>24 hours 

Moderate 

Interference with normal 

activity 

None None A few sensitive 

individuals choose to 

limit activity 

Many sensitive 

individuals choose 

to limit activity 

a An increase in nonspecific bronchial responsiveness of 100% is equivalent to a 50% decrease in PD20 or PD100. 

Source: This table is reproduced from the 1996 O3 AQCD (Table 9-1, page 9-24)

 (U.S. Environmental Protection Agency, 1996a). 

epidemiologically to be associated with ambient O3 exposures in children and adults vigorously 

engaged in outdoor activities. The severity of the symptoms and magnitude of the pulmonary 

decrements, as previously noted, depend on the inhaled O3 dose and individual O3 sensitivity of 

the exposed persons. Controlled exposure chamber studies assessed in the 1996 O3 AQCD 

provided very clear evidence that statistically significant reductions in lung function occurred in 

healthy adults in response to 6.6-h O3 exposures as low as 0.08 ppm under moderate intermittent 

exercise conditions. Of considerable importance, whereas none of the subjects showed marked 

8-67
Table 8-3. Gradation of Individual Responses to Short-Term Ozone Exposure in 

Persons with Impaired Respiratory Systems 

Functional None Small Moderate Large 

FEV1 change Decrements of Decrements of Decrements of Decrements of 

<3% 3 to #10% >10 but <20% $20% 

Nonspecific Within normal Increases of <100% Increases of #300% Increases of >300% 

bronchial range 

responsiveness a 

Airway resistance Within normal SRaw increased <100% SRaw increased up SRaw increased 

(SRaw) range (±20%) to 200% or up to >200% or more than 

15 cm H2O/s 15 cm H2O/s 

Duration of None <4 hours >4 hours but >24 hours 

response #24 hours 

Symptomatic Normal Mild Moderate Severe 

Wheeze None With otherwise normal With shortness of Persistent with 

breathing breath shortness of breath 

Cough Infrequent 

cough 

Cough with deep breath Frequent spontaneous 

cough 

Persistent 

uncontrollable 

cough 

Chest pain None Discomfort just 

noticeable on exercise 

or deep breath 

Marked discomfort 

on exercise or deep 

breath 

Severe discomfort 

on exercise or deep 

breath 

Duration of 

response 

None < 4 hours >4 hours, but 

#24 hours 

>24 hours 

Impact of 

Responses Normal Mild Moderate Severe 

Interference with None Few individuals choose Many individuals Most individuals 

normal activity to limit activity choose to limit choose to limit 

activity activity 

Medical treatment No change Normal medication as Increased frequency Physician or 

needed of medication use or emergency room 

additional medication visit 

a An increase in nonspecific bronchial responsiveness of 100% is equivalent to a 50% decrease in PD20 or PD100. 

Source: This table is reproduced from the 1996 O3 AQCD (Table 9-2, page 9-25) (U.S. Environmental 

Protection Agency, 1996a). 

8-68
decrements in FEV1 measures (e.g., >10% decreases in FEV1) at the end of the 6.6-h exposure to 

filtered air, a notable percentage did so with O3 exposure. More specifically, the percentages 

showing >10% FEV1 decrements at 0.08, 0.10, and 0.12 ppm O3 were 26%, 31%, and 46%, 

respectively, as illustrated in Figure 8-1A. Newly available results from the Adams (2002, 2006) 

studies confirm the finding of a notable percentage (17 to 23%) of healthy adult subjects 

exhibiting >10% decrements in lung function indexed by FEV1 measurements (as shown in 

Figure 8-1B and Figure 8-2). The newly available Adams (2002, 2006) results also further 

extend the earlier findings in terms of providing indications of >10% FEV1 decrements among a 

small percentage (3 to 7%) following 6.6-h exposure of experimentally tested healthy adults to 

0.06 ppm O3 while engaged in moderate exercise (as also shown in Figures 8-1B and 8-2). 

Given that comparable or enhanced pulmonary function decrements are observed in 

asthmatic children (accompanied by respiratory symptoms), it is reasonable to project that at 

least comparable percentages of such children would exhibit >10% FEV1 decrements (and 

respiratory symptoms) at 0.06 or 0.08 ppm if tested under similar exercise conditions. The 

likelihood of notable lung function decrements (as indexed by >10% FEV1 decrease) and/or 

respiratory symptoms occurring among asthmatic children in response to relatively low ambient 

O3 exposures (i.e., <0.08 O3 ppm, 1- to 8-h average) is further supported by the epidemiologic 

observations of the Thurston et al. (1997), Mortimer et al. (2002), and Höppe et al. (2003) 

studies discussed above in Section 8.7.2. The Höppe et al. (2003) study, in fact, suggests that 

about 20% of asthmatic children may be at risk of experiencing >10% FEV1 decrements, which 

presumably could be sufficient to cause many of them to limit activities or increase medication 

use (as indicated in Table 8-3). Such projected potential impacts on asthmatic children in 

response to acute (1- to 8-h) ambient O3 exposures of <0.08 ppm may possibly occur in the range 

of 0.06 to 0.07 ppm, based on the Höppe et al. (2003) study results. The observed lung function 

decrements, as well as any accompanying symptoms and/or increased medication use, would be 

consistent with and lend plausibility to the results of currently available time-series studies that 

found associations between ambient O3 concentrations and increased asthma-related ED visits or 

hospitalizations, especially during the warm season. 

8-69
8.7.5.2 Estimation of Potential Numbers of Persons in At-Risk Susceptible Population 

Groups in the United States 

Although O3-related health risk estimates may appear to be small, they may well be 

significant from an overall public health perspective due to large numbers of individuals in 

potential risk groups. Several subpopulations may be identified as having increased 

susceptibility or vulnerability to adverse health effects from O3, including: older adults, 

children, individuals with preexisting pulmonary disease, and those with higher exposure levels, 

such as outdoor workers. 

One consideration in the assessment of potential public health impacts is the size of various 

population groups that may be at increased risk for health effects associated with O3-related air 

pollution exposure. Table 8-4 summarizes information on the prevalence of chronic respiratory 

conditions in the U.S. population in 2002 and 2003 (Dey and Bloom, 2005; Lethbridge-Çejku 

et al., 2004). Individuals with preexisting cardiopulmonary disease constitute a fairly large 

proportion of the population, with tens of millions of people included in each disease category. 

Of most concern here are those individuals with preexisting respiratory conditions, with 

approximately 11% of U.S. adults and 13% of children having been diagnosed with asthma and 

6% of adults having COPD (chronic bronchitis and/or emphysema). Table 8-5 provides further 

information on the number of various specific respiratory conditions per 100 persons by age 

among the U.S. population during the mid-1990s. Asthma prevalence tends to be higher 

in children than adults. 

In addition, subpopulations based on age group also comprise substantial segments of the 

population that may be potentially at risk for O3-related health impacts. Based on U.S. census 

data from 2003, about 26% of the U.S. population are under 18 years of age and 12% are 

65 years of age or older. Hence, large proportions of the U.S. population are included in age 

groups that are considered likely to have increased susceptibility and vulnerability for health 

effects from ambient O3 exposure. 

The health statistics data illustrate what is known as the “pyramid” of effects. At the top of 

the pyramid, there are approximately 2.5 millions deaths from all causes per year in the U.S. 

population, with about 100,000 deaths from chronic lower respiratory diseases (Kochanek et al., 

2004). For respiratory health diseases, there are nearly 4 million hospital discharges per year 

(DeFrances et al., 2005), 14 million ED visits (McCaig and Burt, 2005), 112 million ambulatory 

8-70
Table 8-4. Prevalence of Selected Respiratory Disorders by Age Group and by Geographic Region in 

(2002 [U.S. Adults] and 2003 [U.S. Children] National Health Interview Survey) 

8-71

Age (Years) R

Adults (18+ Years) 18-44 45-64 65-74 75+ Northeast Mid

Cases 

Chronic Condition/Disease (× 106) % % % % % % % 

Respiratory Conditions 

Asthma 21.9 10.6 11.5 10.6 8.4 7.6 11 10COPD

 Chronic Bronchitis 9.1 4.4 3.5 5.5 5.5 5.3 3.8 

 Emphysema 3.1 1.5 0.3 2 4.9 4.7 1.5 1

Age (Years) R

Children 

(<18 years) 0-4 5-11 12-17 Northeast Mid

Cases 

Chronic Condition/Disease (× 106) % % % % % % Respiratory Conditions 

Asthma 9.1 12.5 7.5 14 14.7 14 13

Source: Lethbridge-Çejku et al. (2004) for data on adults (18+ years); Dey and Bloom (2005) for data on children (<18 years). 
Table 8-5. Acute Respiratory Conditions per 100 Persons/Year by Age Group in the 

United States (1996 National Health Interview Survey) 

45+ Years 

Type of Acute All Under 5 5-17 18-24 25-44 45-64 65+ 

Condition Ages Years Years Years Years Total Years Years 

Respiratory Conditions 78.9 129.4 101.5 86 76.9 53.3 55.9 49 

Common Cold 23.6 48.6 33.8 23.8 18.7 16.1 16.4 15.7 

Other Acute Upper 11.3 13.1 15 16.1 11.6 7 7.5 6.1 

Respiratory Infections 

Influenza 36 53.7 44.3 40.5 38.1 23.3 26.1 18.6 

Acute Bronchitis 4.6 7.2a 4.3 3.9a 5.1 3.8 3.5 4.4a 

Pneumonia 1.8 3.9a 1.7a 1.4a 1.3a 2.0a 0.9a 3.8a 

Other Respiratory 1.7 2.9a 2.4a 0.4a 2.0a 1.1a 1.5a 0.5a 

Conditions 

a Figure does not meet standard of reliability or precision. 

Source: Adams et al. (1999). 

care visits (Woodwell and Cherry, 2004), and an estimated 700 million restricted activity days 

per year due to respiratory conditions (Adams et al., 1999). Combining small risk estimates with 

relatively large baseline levels of health outcomes can result in quite large public health impacts. 

Thus, even a small percentage reduction in O3 health impacts on cardiopulmonary diseases 

would reflect a large number of avoided cases. 

Another key input for public health impact assessment is the range of concentrationresponse functions for various health outcomes. Epidemiologic studies have reported 

associations between short-term exposure to O3 with mortality, hospitalizations for pulmonary 

diseases, ED visits for asthma, reduced lung function, and incidence of respiratory symptoms. 

Effect estimates for morbidity responses to short-term changes in O3 tend to be larger and more 

variable in magnitude than those for mortality. 

In addition to attribution of risks for various health outcomes related to O3 and other 

copollutants, important considerations in assessing the impact of O3 on public health include the 

size of population groups at risk, as well as the concentration-response relationship and potential 

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identification of threshold levels. Taken together, based on the above information, it can be 

concluded that exposure to ambient O3 likely has a significant impact on public health in the 

United States. 

8.8 SUMMARY AND CONCLUSIONS FOR OZONE HEALTH EFFECTS 

This section summarizes the main conclusions derived from this integrated synthesis of 

information regarding health effects associated with ambient O3 exposures. The conclusions are 

based on human clinical, animal toxicologic, and epidemiologic studies that have evaluated 

health effects associated with short-term, repeated, and long-term exposures to O3 alone or in 

combination with other ambient pollutants. The controlled human exposure (or “clinical”) 

studies provide the clearest and most compelling evidence for human health effects directly 

attributable to acute exposures to O3 per se. The evidence from human and animal toxicologic 

studies presented in Chapters 4, 5, and 6 are further useful in not only providing insights into 

possible mechanisms of action underlying different types of O3-related health effects but, also, 

in helping to provide biological plausibility for health effects observed in epidemiologic studies 

assessed in Chapter 7. These empirical efforts are also aimed at identifying susceptible and 

vulnerable populations that are at potentially greater risk for effects of O3 exposure. Overall, the 

new evidence generally supports and builds further upon key health-related conclusions drawn in 

the previous 1996 AQCD. The following conclusions integrate results from newly available 

studies with the scientific evidence assessed in that document. 

1. Health Effects of Short-term Exposures to Ozone 

The 1996 O3 AQCD assessed a substantial body of evidence from toxicologic, human 

clinical, and epidemiologic studies. That AQCD concluded that short-term ambient O3 exposure 

resulted in various respiratory health effects, including lung function decrements and increased 

respiratory symptoms in both healthy and asthmatic individuals exposed during moderate to 

heavy exercise to O3 concentrations ranging down to the lowest levels (0.12 ppm for 1 h; 

0.08 ppm for 6.6 to 8 h) tested in the available controlled human exposure studies. Such 

experimentally demonstrated effects were consistent with and lent plausibility to epidemiologic 

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observations highlighted in the 1996 AQCD of increases in daily hospital admissions and ED 

visits for respiratory causes. Epidemiologic evidence also provided suggestive evidence for an 

association between short-term O3 exposure and mortality. However, there was essentially no 

evidence available in the 1996 O3 AQCD regarding potential cardiovascular effects of short-term 

O3 exposure. The newly-available evidence assessed in this revised O3 AQCD notably enhances 

our understanding of short-term O3 exposure effects, as summarized below, first in relation to 

respiratory morbidity endpoints and then cardiovascular effects and, lastly, mortality. 

A. Respiratory Morbidity 

Lung Function: Controlled exposure studies clearly demonstrate acute reversible 

decrements in lung function in healthy adults exposed to $0.08 ppm O3 when minute ventilation 

and/or duration of exposure are increased sufficiently. On average, spirometric responses to O3 

exposure appear to decline with increasing age starting at approximately 18 years of age. There 

is considerable variability in responses between similarly exposed individuals, such that some 

may experience distinctly larger effects even when small group mean responses are observed. 

For example, healthy adults exposed to 0.08 ppm O3 for 6.6 h with moderate exercise exhibited a 

group mean O3-induced decrement in FEV1 of about 6%, but a decrement of >10% was seen in 

23% of these individuals. Also, exposure to 0.06 ppm O3 caused >10% lung function decline in 

a small percentage (7%) of the subjects. Summer camp field studies conducted in southern 

Ontario, Canada, in the northeastern U.S., and in southern California have also reported lung 

function responses in pre-adolescent children associated with ambient O3 levels. 

With repeated acute O3 exposures (0.12 to 0.45 ppm for 1 h) over several days, controlled 

exposure studies typically find that FEV1 response to O3 is enhanced on the second of several 

days of exposure, but spirometric responses become attenuated on subsequent days with these 

repeated exposures. However, this tolerance is lost after about a week without exposure. 

Animal toxicologic studies also provide extensive evidence that acute O3 exposures alter 

breathing patterns so as to cause rapid shallow breathing (i.e., increased frequency and decreased 

tidal volume), an effect which attenuates after several days of exposure. Such results from 

controlled human exposure studies and animal toxicologic studies provide clear evidence of 

causality for the associations observed between acute (#24 h) O3 exposure and relatively small, 

but statistically significant declines in lung function observed in numerous recent epidemiologic 

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studies. Declines in lung function are particularly noted in children, asthmatics, and adults who 

work or exercise outdoors. 

Respiratory Symptoms: Some young healthy adult subjects exposed in clinical studies to 

O3 concentrations $0.08 ppm for 6 to 8 h during moderate exercise exhibit symptoms of cough 

and pain on deep inspiration. An increase in the incidence of cough has been found in clinical 

studies as low as 0.12 ppm in healthy adults during 1 to 3 h with very heavy exercise and other 

respiratory symptoms, such as pain on deep inspiration and shortness of breath, have been 

observed at 0.16 ppm to 0.18 ppm with heavy and very heavy exercise. These O3-induced 

respiratory symptoms gradually decrease in adults with increasing age. With repeated O3 

exposures over several days, respiratory symptoms become attenuated, but this tolerance is lost 

after about a week without exposure. The epidemiologic evidence shows significant associations 

between acute exposure to ambient O3 and increases in a wide variety of respiratory symptoms 

(e.g., cough, wheeze, production of phlegm, and shortness of breath) in asthmatic children. 

Epidemiologic studies also indicate that acute O3 exposure is likely associated with increased 

asthma medication use in asthmatic children. On the other hand, an effect of acute O3 exposure 

on respiratory symptoms in healthy children is not as clearly indicated by epidemiology studies, 

consistent with diminished symptom responses seen in healthy children in human clinical 

studies. 

Airway Inflammation: Inflammatory responses have been observed subsequent to 6.6 h O3 

exposures to the lowest tested level of 0.08 ppm in healthy human adults. Some studies suggest 

that inflammatory responses may be detected in some individuals following O3 exposures even 

in the absence of O3-induced pulmonary function decrements in those subjects. With repeated 

O3 exposures over several days, an attenuation of most inflammatory markers occurs. However, 

none of the several markers of lung injury and permeability evaluated show attenuation, 

indicating continued lung tissue damage during repeated exposure. Animal toxicologic studies 

provide extensive evidence that acute (1 to 3 h) O3 exposures as low as 0.1 to 0.5 ppm can cause 

(1) lung inflammatory responses (typified by increased reactive oxygen species, inflammatory 

cytokines, influx of PMNs, and activation of alveolar macrophages); (2) damage to epithelial 

airway tissues, (3) increases in permeability of both lung endothelium and epithelium, and 

(4) increases in susceptibility to infectious diseases due to modulation of lung host defenses. 

Consistent with these experimental findings, there is also limited epidemiologic evidence 

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showing an association between acute ambient O3 exposure and airway inflammation in children 

acutely exposed to ambient O3 concentrations (1-h max O3 of approximately 0.1 ppm). The 

extensive human clinical and animal toxicological evidence, together with the limited available 

epidemiologic evidence, is clearly indicative of a causal role for O3 in inflammatory responses in 

the airways. 

Airway Responsiveness: Controlled human exposure studies have found that acute O3 

exposure causes an increase in nonspecific airway responsiveness, as indicated by reductions in 

concentrations of methacholine or histamine required to produce a given decrease in FEV1 or 

increase in SRaw. Acute (2- or 3-h) O3 exposure at 0.25 or 0.4 ppm of allergic asthmatic subjects, 

who characteristically already have somewhat increased airway responsiveness at baseline, was 

found to cause further increases in airway responsiveness in response to allergen challenges. 

Also, repeated daily exposure to 0.125 ppm O3 for 4 days exacerbated lung function decrements 

in response to bronchial allergen challenges among persons with preexisting allergic airway 

disease, with or without asthma. Ozone-induced exacerbation of airway responsiveness persists 

longer and attenuates more slowly than O3-induced pulmonary function decrements and 

respiratory symptom responses. Heightened airway responsiveness (reactivity) has also been 

observed in several laboratory animal species with acute exposures (1 to 3 h) to 0.5 to 1.0 ppm 

O3. Ozone increases airway hyperreactivity to bronchoconstrictive agents (e.g., ovalbumin), and 

there is a temporal relationship between inflammatory cell influx and O3-induced increases in 

airway reactivity. Several studies of sensitized laboratory animals showing O3-induced increases 

in airway hyperreactivity are consistent with O3 exacerbation of airway hyperresponsiveness 

reported in atopic humans with asthma. Airway responsiveness has not been widely examined in 

epidemiologic studies. However, the evidence from human clinical and animal toxicological 

studies clearly indicate that acute exposure to O3 can induce airway hyperreactivity, thus likely 

placing atopic asthmatics at greater risk for more prolonged bouts of breathing difficulties due to 

airway constriction in response to various airborne allergens or other triggering stimuli. 

Respiratory Hospital Admissions and Emergency Department Visits: Aggregate 

population time-series studies observed that ambient O3 concentrations are positively and 

robustly associated with respiratory-related hospitalizations and asthma ED visits during the 

warm season. These observations are strongly supported by the human clinical, animal 

toxicologic, and epidemiologic evidence for lung function decrements, increased respiratory 

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symptoms, airway inflammation, and airway hyperreactivity. Taken together, the overall 

evidence supports a causal relationship between acute ambient O3 exposures and increased 

respiratory morbidity outcomes resulting in increased ED visits and hospitalizations during the 

warm season. 

B. Cardiovascular Morbidity 

At the time of the 1996 O3 AQCD, the possibility of O3-induced cardiovascular effects was 

a largely unrecognized issue. Newly-available evidence has emerged since then which provides 

considerable plausibility for how O3 exposure could exert cardiovascular impacts. This includes 

direct O3 effects such as O3-induced release from lung epithelial cells of platelet activating factor 

(PAF) that may contribute to blood clot formation that would increase the risk of serious 

cardiovascular outcomes (e..g, heart attack, stroke, mortality). Also, interactions of O3 with 

surfactant components in epithelial lining fluid of the lung results in production of oxysterols 

and reactive oxygen species that may exhibit PAF-like activity contributing to clotting and/or 

exert cytotoxic effects on lung and heart cells. Other possible mechanisms may involve O3­

induced secretions of vasoconstrictive substances and/or effects on neuronal reflexes that may 

result in increased arterial blood pressure and/or altered electrophysiologic control of heart rate 

or rhythm. Consistent with the latter possibility, some field/panel studies that examined 

associations between O3 and various cardiac physiologic endpoints have yielded limited 

epidemiologic evidence suggestive of a potential association between acute O3 exposure and 

altered HRV, ventricular arrhythmias, and incidence of MI. Also, highly suggestive evidence for 

O3-induced cardiovascular effects is provided by a few population studies of cardiovascular 

hospital admissions which reported positive O3 associations during the warm season between 

ambient O3 concentrations and cardiovascular hospitalizations. On the other hand, the only 

controlled human exposure study that evaluated effects of O3 exposure on cardiovascular health 

outcomes found no significant O3-induced differences in ECG, heart rate, or blood pressure in 

healthy or hypertensive subjects, but did observe an overall increase in myocardial work and 

impairment in pulmonary gas exchange. Also, some animal toxicological studies have shown 

O3-induced decreases in heart rate, mean arterial pressure, and core temperature. Overall, then, 

this generally limited body of evidence is highly suggestive that O3 directly and/or indirectly 

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contributes to cardiovascular-related morbidity, but much remains to be done to more fully 

substantiate links between ambient O3 exposure and adverse cardiovascular outcomes. 

C. Mortality 

Numerous recent epidemiologic studies conducted in the United States and abroad have 

investigated the association between acute exposure to O3 and mortality. Results from several 

large U.S. multicity studies as well as several single-city studies indicate a positive association 

between increases in ambient O3 levels and excess risk of all-cause (nonaccidental) daily 

mortality. Determining cause-specific mortality is more difficult due to reduced statistical 

power by which to examine cause-specific associations and the lack of clarifying information on 

contributing causes of death. That is, attribution to one or the other of the more specific 

cardiopulmonary causes may underplay contributions of chronic cardiovascular disease to 

“respiratory” deaths (e.g., a heart attack victim succumbing to acute pneumonia) or vice versa. 

Nevertheless, consistent with observed O3-related increases in respiratory- and cardiovascularrelated morbidity, several newer multicity studies, single-city studies, and several meta-analyses 

of these studies have provided relatively strong epidemiologic evidence for associations between 

short-term O3 exposure and all-cause mortality, even after adjustment for the influence of season 

and PM. In addition, consistently positive associations have been reported for O3-related 

cardiovascular mortality across approximately 30 studies, with two well-conducted multicity 

studies in the United States and Europe yielding small, but statistically significant positive 

associations. Also, as discussed in Section 8.6, newly available experimental data from both 

animal and human studies provide evidence suggestive of plausible pathways by which risk of 

respiratory or cardiovascular morbidity and mortality could be increased by ambient O3 either 

acting alone or in combination with copollutants in ambient air mixes. This overall body of 

evidence is highly suggestive that O3 directly or indirectly contributes to non-accidental and 

cardiopulmonary-related mortality, but additional research is needed to more fully establish 

underlying mechanisms by which such effects occur. 

2. Health Effects of Long-term Exposures to Ozone 

In the 1996 O3 AQCD, the available epidemiologic data provided only suggestive evidence 

that respiratory health effects were associated with chronic O3 exposure. Animal toxicologic 

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studies indicated that chronic O3 exposure caused structural changes in the respiratory tract, and 

simulated seasonal exposure studies in animals suggested that such exposures might have 

cumulative impacts. As summarized below, recent studies are generally consistent with the 

conclusions drawn in the previous 1996 AQCD. 

A. Respiratory Morbidity 

Lung Function: Recent epidemiologic studies observed that reduced lung function growth 

in children was associated with seasonal exposure to O3; however, cohort studies investigating 

the effect of annual or multiyear O3 exposure observed little clear evidence for impacts of 

longer-term, relatively low-level O3 exposure on lung function development in children. The 

epidemiologic data, collectively, indicate that the current evidence is suggestive but inconclusive 

for respiratory health effects from long-term O3 exposure. 

Morphological Changes: Animal toxicologic studies continue to show chronic O3-induced 

structural alterations in several regions of the respiratory tract including the centracinar region. 

Morphologic evidence from some recent studies using exposure regimens that mimic seasonal 

exposure patterns report increased lung injury compared to conventional chronic stable 

exposures. Infant rhesus monkeys repeatedly exposed to 0.5 ppm 8h/day O3 for 11 episodes 

exhibited: (1) remodeling of the distal airways; (2) abnormalities in tracheal basement 

membrane; (3) eosinophil accumulation in conducting airways; and (4) decrements in airway 

innervation. Long-term O3 exposure of rats to 0.5 or 1.0 ppm for 20 months resulted in upper 

respiratory tract mucus metaplasia and hyperplasia in the nasal epithelium (0.25 or 0.5 ppm, 

8h/day, 7days/wk for 13 weeks). The persistent nature of these cytological changes raise the 

possibility of long-lasting alterations in human airways in response to chronic O3 exposure, but it 

is highly uncertain as to what long-term patterns of exposure or O3 concentrations in humans 

may be requisite to produce analogous morphological changes. Nor is it now possible to 

characterize the possible magnitude or severity of any such effects occurring in humans in 

response to ambient O3 exposures at levels observed in the United States. 

Incidence of Lung Cancer: The weight of evidence from recent animal toxicological 

studies and a very limited number of epidemiologic studies do not support ambient O3 as a 

pulmonary carcinogen. 

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B. Mortality 

Results from the few available epidemiologic studies are inconsistent regarding the 

association between long-term exposure to O3 and mortality. There is little evidence to suggest a 

causal relationship between chronic O3 exposure and increased risk for mortality in humans. 

3. Susceptibility or Vulnerability to Effects Associated with Exposure to Ozone 

Various factors have been shown to influence individuals’ responses to environmental air 

pollutants. Factors that increase susceptibility to O3-related effects include innate factors, such 

as genetic predisposition or developmental effects, or disease status. Other factors can lead to 

enhanced vulnerability to O3-related effects, such as heightened exposures or activity patterns. 

In the 1996 O3 AQCD, available evidence suggested that children, asthmatics, and outdoor 

workers were populations that may be more susceptible or vulnerable to effects of O3 exposure. 

In addition, controlled human exposure studies also demonstrated a large variation in sensitivity 

and responsiveness to O3 in studies of healthy subjects, but the specific factors that contributed 

to this intersubject variability were yet to be identified. Recent studies have built upon the 

evidence available in the previous review. Factors related to susceptibility or vulnerability to O3 

exposure-related effects are briefly summarized below: 

People with Preexisting Pulmonary Diseases: Ozone-induced differential responses in 

lung function and AHR in people with allergic rhinitis suggest that asthmatics have potentially 

greater responses than healthy people with exposure to O3. There is a tendency for slightly 

increased spirometric responses in mild asthmatics and allergic rhinitics relative to healthy 

young adults. Spirometric responses in asthmatics appear to be affected by baseline lung 

function, i.e., responses increase with disease severity. In addition, repeated O3 exposure over 

several days has been shown to increase responsiveness to bronchial allergen challenge in 

subjects with preexisting allergic airway disease, with or without asthma. Asthmatics also show 

a significantly greater neutrophil response (18 h postexposure) than similarly-exposed healthy 

individuals. Epidemiologic studies have reported associations with a range of respiratory health 

outcomes in asthmatics, from decreases in lung function to hospitalization or ED visits for 

asthma, thus supporting this population group as being likely to experience increased risk for O3­

induced health effects. Although controlled human exposure studies have not found evidence of 

larger spirometric changes in people with COPD relative to healthy subjects, this may be due to 

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the fact that most people with COPD are older adults who would not be expected to have such 

changes based on their age. However, new epidemiologic evidence indicates that people with 

COPD may be more likely to experience other effects, including emergency room visits, hospital 

admissions, or premature mortality. 

Age-related: Controlled human exposure studies have shown that lung function responses 

to O3 varies with age, with responsiveness generally diminishing after about 18 to 20 years of 

age. Children and older adults thus have lesser respiratory symptoms with O3 exposure than 

young healthy adults. Potentially increased O3 doses can be received by individuals 

experiencing less severe respiratory symptoms. Evidence from newer epidemiologic studies 

supports the 1996 O3 AQCD conclusions that children are more likely at increased risk for O3­

induced health effects. Notably, epidemiologic studies have indicated adverse respiratory health 

outcomes associated with O3 exposure in children. In addition, recently published epidemiologic 

studies also suggest that older adults (aged $65 years) appear to be at excess risk of O3-related 

mortality or hospitalization. 

Heightened vulnerability due to greater exposures: Epidemiologic studies have provided 

some evidence to indicate that outdoor workers are more vulnerable to O3-related effects, which 

is likely related to their increased exposure to ambient air pollution. Controlled human exposure 

studies clearly established differential biological response to O3 based on physical activity 

(exertion). Epidemiologic studies also suggest that exercising (moderate to high physical 

exertion) children and adolescents appear to demonstrate increased responsiveness to ambient 

concentrations of O3 and may be more likely to experience O3-induced health effects. Animal 

studies show a similar impact of exercise on responsiveness to O3. 

Genetic susceptibility: Animal toxicologic studies provide supportive evidence to the 

observations of innate susceptibility. Various strains of mice and rats have demonstrated the 

importance, in general, of genetic background in O3 susceptibility. Moreover, genetic and 

molecular characterization studies in laboratory animals identified genetic loci responsible for 

both sensitivity and resistance. Recent human clinical and epidemiologic studies also have 

shown that genetic polymorphisms for antioxidant enzymes and inflammatory genes (GSTM1, 

NQO1, and Tmf-") may modulate the effect of O3 exposure on pulmonary function and airway 

inflammation. 

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4. Health Effects of Ozone-Containing Pollutant Mixtures 

The potential interaction of pollutant mixtures with O3 is poorly understood and the animal 

studies reviewed in the 1996 O3 AQCD reported additive, synergistic or even antagonistic effects 

depending on the exposure regimen and the endpoint studied. A few new controlled human 

exposure and animal toxicology studies reviewed in Chapters 4, 5, and 6 investigated health 

effects associated with O3-containing pollutant mixtures of near ambient levels. As noted below, 

recent studies, although generally consistent with conclusions drawn in the 1996 O3 AQCD, have 

added some new information, particularly with respect to interactions between O3 and PM. 

Controlled human exposure studies indicate that continuous exposure of healthy human 

adults to SO2 or NO2 increases bolus dose O3 absorption, suggesting that co-exposure to other 

gaseous pollutants in the ambient air may enhance O3 absorption. Other controlled human 

exposure studies that evaluated response to allergens in asthmatics (allergic and dust-mite 

sensitive) suggest that O3 enhances response to allergen challenge. Consistent with these 

findings, animal toxicology studies also reported enhanced response to allergen on exposure 

to O3. 

A few other animal toxicology studies that exclusively investigated the co-exposure of PM 

and O3 reported increased response (lung tissue injury, inflammatory and phagocytosis) to the 

mixture of PM + O3 compared to either PM or O3 alone. Recent investigations on the 

copollutant interactions using simulated urban photochemical oxidant mixes suggest the need for 

similar studies in understanding the biological basis for air pollutant mixture effects observed in 

epidemiologic studies. 

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8-92
APPENDIX 8A

Summary of New Animal Toxicology, 

Human Clinical, and U.S./Canadian Epidemiologic Studies 

of Health Effects Associated with Ambient or 

Near-Ambient Ozone Exposures 

8A-1
Table 8A-1. Short-Term Ozone-Induced Health Effects Observed in Controlled Human Expos

Health Ozone Exposure Subject Effects Concentration Duration/Activity Characteristics Observable Effects 

Pulmonary Function Short-term (1-2 h) 

0.0!0.4 2 h rest or IE 

(4 × 15 min at V

. 

E = 25 or 35 

L/min/m2 BSA) 

Healthy NS 18 to 36 yr old mean age 24 yr 

Statistical analysis of 8 experimental chastudies conducted between 1980 and 199U.S. EPA in Chapel Hill, NC. Response decreased with age, was minimally affecbody size corrections, and was not more to O3 concentration than V. E. 

Prolonged (6.6 h) 

0.0!0.08 6.6 h 

IE (6 × 50min) V

. 

E = 20 L/min/m2 BSA 

Healthy NS Males 23.5 ± 3.0 yrs Females 22.8 ± 1.2 yrs 

FEV1 and symptom responses after 6.6 h 

exposure to 0.04 and 0.06 ppm not signidifferent from FA. 

0.0!0.12 6.6 h 

IE (6 × 50min) V

. 

E = 20 L/min/m2 BSA 

Healthy NS, 22.4 ± 2.4 yrs old 

FEV1 and total symptoms after 6.6 h exp

0.04 ppm not significantly different from 

(a) 0.08 6.6 h 

IE (6 × 50 min) V

. 

E = 20 L/min/m2 BSA 

Healthy NS, 18 to 25 yr old 

(a) FEV1 decreased 6.2% after 6.6 h in s

wave exposures. Total symptoms signifincreased at 5.6 and 6.6 h. 

(b) 0.08 (mean) varied from 0.03 to 0.15 

(b) FEV1 decreased 5.6 to 6.2% after 4.6 

respectively, in varied exposure; total sysignificantly increased also after 4.6 to 6

8A-2

Repeated 0.12 6.6 h Healthy NS FEV1 responses were maximal on first d

50 min exercise/10 min exposure (!13%), less on second day (!rest, 30 min lunch V

. 

E = 38.8 L/min 

absent thereafter. Symptoms only the firMethacholine airway responsiveness wadoubled on all exposure d, but was highesecond day of O3. Airway responsivene

still higher than air control after 5 d of O

exposure. Trend to lessened response, bnot achieved after 5 d. 
Table 8A-1 (cont’d). Short-Term Ozone-Induced Health Effects Observed in Controlled Human Ex

Health Ozone Exposure Subject Effects Concentration Duration/Activity Characteristics Observable Effects 

Pulmonary Function (cont’d) Preexisting Disease 

0.125 to 0.250 

0.125 

3 h 

IE (10 min rest, 15 min exercise on bicycle) V

. 

E = 30 L/min 

3 h IE × 4 d 

Mild bronchial asthma 20-53 yr old Allergic rhinitis 19-48 yr old 

Mean early-phase FEV1 response and nu

$20% reductions in FEV1 were significa

greater after 0.25 ppm O3 or 4 × 0.125 pp

Most of the $15% late-phase FEV1 respo

occurred after 4 d of exposure to 0.125 p

as well as significant inflammatory effecindicated by increased sputum eosinophi(asthma and allergic rhinitis) and increassputum lymphocytes, mast cell tryptase, histamine, and LDH (asthma only). 

0.0 0.12 

0.75 h IE (15 min exercise, 15 min rest) V

. 

E = 40-46 L/min 

Asthmatics sensitive to SO2 19 to 38 yrs 

old 

No significant differences due to O3 betw

placebo and antioxidant supplement coheither spirometric responses or bronchial hyperresponsiveness to 0.1 ppm SO2. 

Copollutants 

SO2 0.10 to 0.12 1 h (mouthpiece) 

IE V

. 

E . 30 L/min 

45-min exposure to air or O3, followed by 

15-min exposure to O3 

or SO2 

Allergic asthmatics, 12 to 18 yr old, medications withheld for at least 4 h before exposures 

Prior exposure to O3 potentiated pulmon

function responses to SO2; decrements in 

NO2 0.0 to 0.3 1 h (mouthpiece) 

IE V

. 

E = 33 L/min 

V

. 

E = 35 L/min 

Healthy NS, 12 to 17 yr old Asthmatic 13 to 18 yr old 

No significant differences in FEV1 and R

between asthmatics and healthy, or betwatmospheres and cohorts. 

8A-3
Table 8A-1 (cont’d). Short-Term Ozone-Induced Health Effects Observed in Controlled Human Ex

8A-4

Health Ozone Exposure Subject Effects Concentration Duration/Activity Characteristics Observable Effects 

Airway Hyperresponsiveness Preexisting 0.125 3 h Mild bronchial Mean early-phase FEV1 response and nuDisease 0.250 IE (10 min rest, 15 min asthma of $20% reductions in FEV1 were signif

exercise on bicycle) V

. 

E = 30 L/min Allergic rhinitis 

greater after 0.25 ppm O3 or 4 × 0.125 pp

Most of the $15% late-phase FEV1 respo

3h IE × 4 d occurred after exposure to 4 × 0.125 ppm 

as well as significant inflammatory effecindicated by increased sputum eosinophi(asthma and allergic rhinitis) and increassputum lymphocytes, mast cell tryptase, histamine, and LDH (asthma only). 

0.12 45 min Physician diagnosed The authors concluded O3 exposure incr

IE (exercise, rest, asthma; SO2 - bronchial responsiveness to SO2 in asthmexercise) .

VE = 3 × resting 

induced airway hyperreactivity 

and that antioxidant supplementation has protective effects against this responsiveespecially in the “more-severe” responde

0.12 1 h rest Mild allergic No effect of O3 on airway response to grAir antigen asthma; ragweed allergen. 

18 to 49 yr of age 

0.12 Rest Atopic No effect of O3 on airway response to gr

asthma allergen. 

0.10 to 0.40 1 h Stable mild No significant differences in FEV1 or FV

Light IE (2 × 15 min on asthmatics with observed for 0.10 and 0.25 ppm O3-FA etreadmill) .

VE = 27 L/min 

FEV1 >70% and 

methacholine 

or postexposure exercise challenge; 12 sexposed to 0.40 ppm O3 showed signific

responsiveness reduction in FEV1. 

Healthy 0.08 to 0.12 6.6 h Healthy NS, 33, 47, and 55% decreases in cumulative Subjects IE at .39 L/min 18 to 32 yr old methacholine required to produce a 100% 

increase in SRaw after exposure to O3 at 

0.10, and 0.12 ppm, respectively. 
Table 8A-1 (cont’d). Short-Term Ozone-Induced Health Effects Observed in Controlled Human Ex

Health Ozone Exposure Subject 

Effects Concentration Duration/Activity Characteristics Observable Effects 

Pulmonary Inflammation Short-term (1-2 h) 

0.1 2 h 

Mild IE 

Healthy subjects mean age ~30 yr 

Markers of exposure in exhaled breath condensate, including increased 8-isoproTBARS and LTB-4, and a marker of ROinteraction in peripheral blood leukocyte(8-OHdG), were increased in a sub-set osubjects bearing the wild genotype for NAD(P)H:quinone oxidoreductase and tgenotype for glutathione-S- transferase M

0.12 2 h 

IE (15 min/30 min); ( 

.

VE) . 20 L/min/m2 

BSA 

Healthy nonsmokers; mean age ~28 yr 

Increase in the percentage of vessels expP-selectin in bronchial biopsies at 1.5 h postexposure. No changes in FEV1, FVC

inflammatory cells or markers in BAL, oexpressing VCAM-1, E-selectin or ICAMbiopsies. 

8A-5

Preexisting Disease 

0.125 0.25 

3 h exposures to both O3 

concentrations and to FA; 3 h on four consecutive days to 0.125; study arms separated by >4 wks IE (15 min/30 min) 

Allergic asthmatic and allergic rhinitic subjects; 19-53 yr of age 

Repeated exposure caused increases in nand eosinophil numbers in both subject gas well as increased percentage and numlymphocytes in the asthmatics. 

Cardiovascular Copollutants 

PM2.5 0.0 to 0.12 2-2.5 h Healthy NS Neither systolic nor diastolic pressure has be

rest 18 to 50 yr old affected by pollutants exposure despite a sign

brachial artery constriction and a reduction in diameter when compared to filtered air (p = 0
Table 8A-2. Effects of Acute O3 Exposure on Lung Function in the U.S. and Cana

8A-6

Statistics for 8-h max O3 

Air Quality Data (ppb) a Standardized Percent Change 

Reference, Study Mean O3 Levels 

Location and Period Study Population (ppb) 98th % 99th % Range Morning After

Brauer et al. (1996) Berry pickers aged 1-h max O3: 55 55 3-55 FEV1: FEV1: Fraser Valley, British 10-69 yr (n = 58) Columbia, Canada repeatedly monitored 40.3 Lag 1: Lag 0: Jun-Aug 1993 over a 59-d period. SD 15.2 –6.4% (–8.0, –4.7) –5.4% (–6.5

Outdoor work shifts averaged 11 h in duration. 

Mortimer et al. National Cooperative 8-h avg O3 64.3 66 28.8-66 PEF: — (2000; 2002) Inner City Asthma (10 a.m.-6 p.m.): Eight urban areas in Study (NCICAS) Lag 1-5: the U.S.: Baltimore, cohort. Asthmatic 48 All areas: MD; Bronx, NY; children aged 4-9 yr SD not provided. –1.18% (–2.10, –0.26) Chicago, IL; (n = 846) repeatedly Cleveland, OH; monitored over a Range of Detroit, MI; 3-mo period. medians across East Harlem, NY; cities: St. Louis, MO; Approximately Washington, DC 34 to 58 Jun-Aug 1993 (<5% of days 

exceeded 80). 

Ross et al. (2002) Mild and severe 8-h max O3: 68.8 75 8.9-78.3 PEF: PEF: East Moline, IL and asthmatics aged 5-49 nearby communities yr (n = 40) repeatedly 41.5 Lag 0-1: Lag 0: May-Oct 1994 monitored for a 5-mo SD 14.2 !0.96% (!1.78, !0.14) !1.08% (!1

period. IQR 20 

Naeher et al. (1999) Nonsmoking women 24-h avg O3: 74 79 13-87 PEF: PEF: Vinton, VA aged 19-43 yr Summers 1995, 1996 (n = 473) who 34.87 Lag 1: Lag 0: 

recently delivered SD 8.86 !0.31% (!0.68, 0.07) !0.36% (!0babies repeatedly Range 8.74-56.63 monitored for a Lag 1-3: Lag 1-5: two-wk period. !0.52% (!1.11, 0.07) !1.11% (!1
Table 8A-2 (cont’d). Effects of Acute O3 Exposure on Lung Function in the U.S. and C

8A-7

Statistics for 8-h max O3 

Air Quality Data (ppb) a Standardized Percent Change 

Reference, Study Mean O3 Levels 

Location and Period Study Population (ppb) 98th % 99th % Range Morning After

Korrick et al. (1998) Hikers aged 15-64 yr Avg of hourly O3 87 89 24-91 — — Mount Washington, (n = 530) monitored during each hike NH before and after their (approximately Summers 1991, 1992 hike. Hikes averaged 8-h avg O3): 

8 h in duration. 

40 SD 12 Range 21-74 

Neas et al. (1995) Symptomatic and All 12-h avg O3 96.5 98 15-98 — PEF: Uniontown, PA asymptomatic 4th and (8 a.m.-8 p.m. Summer 1990 5th grade children and 8 p.m.-8 Daytime 12-

(n = 83) who did not a.m.): Lag 0: use any asthma !0.62% (!1medication during the 37.2 previous year. Each IQR 29.9 Weighted by child, on average, was Maximum 87.5 proportion omonitored on 43 d. spent outdoo

Daytime prior 12 h: 12-h avg O3 Lag 0: (8 a.m.-8 p.m.): !0.78% (!150.0 SD not provided. 
Table 8A-2 (cont’d). Effects of Acute O3 Exposure on Lung Function in the U.S. and C

8A-8

Statistics for 8-h max O3 

Air Quality Data (ppb) a Standardized Percent Change 

Reference, Study Mean O3 Levels 

Location and Period Study Population (ppb) 98th % 99th % Range Morning After

Neas et al. (1999) Children aged 6-11 yr 12-h avg O3 96.9 104.5 17.7-104.5 PEF: PEF: Philadelphia, PA (n = 156) at two (9 a.m.-9 p.m.): Jul-Sep 1993 summer camps Lag 1: Lag 0: 

repeatedly monitored SW camp: !0.74% (!1.54, 0.07) !0.46% (!1over 40 d. 57.5 

IQR 19.8 Lag 1-5: Lag 1-5: 

!0.76% (!2.65, 1.13) !0.26% (!1

NE camp: 55.9 IQR 21.9 

Delfino et al. (1997) Symptomatic 12-h avg O3 110 121 38-121 PEF: PEF: Alpine, CA asthmatics, children (8 a.m.-8 p.m.): May-Jul 1994 aged 10 to 15 yr No effects observed No effects o

(n = 13) and adults Ambient: using ambient or using ambieaged 24 to 47 yr 64 personal O3 personal O3 (n = 9), repeatedly SD 17 concentrations. concentratiomonitored for a Range 34-103 8-week period. Effect estimates not Effect estim

Personal: presented. presented. 18 SD 14 Range 0-80 55% of personal O3 samples 

below limit of detection. 

Linn et al. (1996) School children 24-h avg O3: 150 164 2.5-192.5 FEV1: FEV1: Three towns in (n = 269) repeatedly California: monitored for one 23 Lag 1: Lag 0: Rubidoux, Upland, week in fall, winter, SD 12 –0.27% (–0.79, 0.24) –0.19% (–0.Torrance and spring during their Range 3-53 Fall-spring 1992­ 4th and 5th grade 1993 and 1993-1994 school years. 
Table 8A-2 (cont’d). Effects of Acute O3 Exposure on Lung Function in the U.S. and C

8A-9

Statistics for 8-h max O3 

Air Quality Data (ppb) a Standardized Percent Change 

Reference, Study Mean O3 Levels 

Location and Period Study Population (ppb) 98th % 99th % Range Morning After

Thurston et al. (1997) Children aged 7-13 yr 1-h max O3: — — — — — Connecticut River (total n = 166) with Valley, CT moderate-to-severe 1991: 114.0 June 1991, 1992, asthma repeatedly 1992: 52.2 1993 monitored over a 5-d 1993: 84.6 

period. 

1991-1993: 83.6 

a Using O3 data obtained for the study period in the location of the study, 8-h max O3 concentrations were derived and statistics were calculated. The 

 for the full study period distribution are presented here (unless noted otherwise), along with the range (minimum-maximum) of concentrations. Sin

 vary in length, from several weeks to over 10 yr, the 98th and 99th percentile values were selected for presentation here as a high study period conc

 approximates a 4th maximum concentration, depending on the study period length. 

b Percent change in lung function per standard unit ppb O3: 40 ppb for 1-h max O3; 30 ppb for 8-h max O3 or 8-h avg O3; 25 ppb for 12-h avg O3; an
Table 8A-3. Effects of Acute O3 Exposure on Asthma Emergency Department Visits in the U.S. 

Statistics for 8-h max O3 

Air Quality Data (ppb) a Standardized Percent Exc

Reference, Study Mean O3 Levels 

Location and Period (ppb) 98th % 99th % Range All Year 

8A-10

Stieb et al. (1996) Saint John, New Brunswick, Canada May-Sep 1984-1992 

1-h max O3: 

Warm season: 41.6 Range 0-160 95th % 75 

83 91 5-140.5 — 

Cassino et al. (1999) New York City 1992-1995 

24-h avg O3: 

All yr: 17.5 IQR 14 

83.3 88.8 3-114.6 All ages: 

All subjects: Lag 2: 8.3% (!0.8, 18.5) Note: Used Poisson GAM with default convergence criteria 

Wilson et al. (2005) Manchester, NH 1996-2000 

8-h max O3: 

Spring (Mar-May): 43.4 SD 9.7 

Apr-Oct: 85 Apr-Oct: 93 Apr-Oct: 

5-121 

— 

Summer (Jun-Aug): 42.8 SD 14.6 

Friedman et al. (2001) Atlanta, GA Jun-Sep 1996 

1-h max O3: 

Baseline (Jun 21-Jul 18 and Aug 5-Sep 1): 81.3 SD not provided. 

Jul 19-Aug 

4: 85.8 

Jul 19-Aug 

4: 85.8 

Jul 19-Aug 

4: 20-85.8 

— 

Intervention period (Jul 19-Aug 4): 58.6 SD not provided. 
Table 8A-3 (cont’d). Effects of Acute O3 Exposure on Asthma Emergency Department Visits in the U

Statistics for 8-h max O3 

Air Quality Data (ppb) a Standardized Percent Exc

Reference, Study Mean O3 Levels 

Location and Period (ppb) 98th % 99th % Range All Year 

Jaffe et al. (2003) 8-h max O3: 98 106 25-117 — Columbus, OH Jun-Aug 1991-1996 Summer: 57 

SD 16 

Jaffe et al. (2003) 8-h max O3: 104 107 27-111 — Cleveland, OH Jun-Aug 1991-1996 Summer: 50 

SD 17 

Jaffe et al. (2003) 8-h max O3: 104 108 24-124 — Cincinnati, Cleveland, and Columbus, OH Summer: Jun-Aug 1991-1996 All three cities: 

Not reported. 

Jaffe et al. (2003) 8-h max O3: 106 116 24-124 — Cincinnati, OH Jun-Aug 1991-1996 Summer: 60 

SD 20 

8A-11
Table 8A-3 (cont’d). Effects of Acute O3 Exposure on Asthma Emergency Department Visits in the U

Statistics for 8-h max O3 

Air Quality Data (ppb) a Standardized Percent Exc

Reference, Study 

Location and Period 

Mean O3 Levels 

(ppb) 98th % 99th % Range All Year 

Wilson et al. (2005) Portland, ME 1998-2000 

8-h max O3: 

Spring (Mar-May): 43.7 SD 10.2 

Apr-Oct: 

108 

Apr-Oct: 121 Apr-Oct: 

15-142 

— 

Summer (Jun-Aug): 46.1 SD 15.4 

Tolbert et al. (2000) Atlanta, GA Jun-Aug 1993-1995 

8-h max O3: 

Summer: 59.3 SD 19.1 

108.9 112.6 16.2-135.8 — 

Peel et al. (2005) Atlanta, GA 1993-2000 

8-h max O3: 

Mar-Nov: 55.6 SD 23.8 

Mar-Nov: 

115 

Mar-Nov: 

124 

Mar-Nov: 

3-152 

All available data (Mar-Nov): All ages: Lag 0-2: 2.6% (!0.5, 5.9) 

8A-12

a Using O3 data obtained for the study period in the location of the study, 8-h max O3 concentrations were derived and statistics were calculate

 percentile values for the full study period distribution are presented here (unless noted otherwise), along with the range (minimum-maximum

 Since the time periods of the studies vary in length, from several weeks to over 10 yr, the 98th and 99th percentile values were selected for p

 high study period concentration that roughly approximates a 4th maximum concentration, depending on the study period length. 

b Percent change in lung function per standard unit ppb O3: 40 ppb for 1-h max O3; 30 ppb for 8-h max O3; and 20 ppb for 24-h avg O3. 
Table 8A-4. Effects of Acute O3 Exposure on Total Respiratory and Asthma Hospital Adm

in the U.S. and Canada 

Statistics for 8-h max O3 

Air Quality Data (ppb) a Standardized Percent Exc

Reference, Study 

Location and Period Mean O3 Levels (ppb) 98th % 99th % Range All Year 

8A-13

Yang et al. (2003) 24-h avg O3: 42.7 47.3 1.1-71.9 Total respiratory: Vancouver, British Columbia, Canada All yr: 13.41 Age <3 yr: 1986-1998 SD 6.61 Lag 4: 50.4% (33.2, 71.4) 

IQR 9.74 

Age 65+ yr: Lag 4: 28.5% (19.4, 40.5) 

Burnett et al. (1997a) 1-h max O3: Apr-Dec: Apr-Dec: Apr-Dec: — 16 Canadian cities 47.1 51.3 6.2-68.4 1981-1991 All yr: 31 

95th % 60 Mean range across cities: 26-38 95th % 45-84 Apr-Dec: 32.9 95th % 64 

Burnett et al. (1997b) 1-h max O3: 62 64 0-79 — Toronto, Ontario, Canada Summers 1992-1994 Summer: 41.2 

IQR 22 12-h avg O3: 

Summer: IQR 11.5 

Sheppard et al. 8-h max O3: 65 73 2-100 Asthma: (1999; reanalysis Sheppard, 2003) All yr: 30.4 Age 0-64 yr: Seattle, WA IQR 20 Lag 2: 10.7% (1.5, 20.1) 1987-1994 
Table 8A-4 (cont’d). Effects of Acute O3 Exposure on Total Respiratory and Asthma Hospital A

in the U.S. and Canada 

Statistics for 8-h max O3 

Air Quality Data (ppb) a Standardized Percent Exc

Reference, Study 

Location and Period Mean O3 Levels (ppb) 98th % 99th % Range All Year 

8A-14

Lin et al. (2004) Vancouver, British Columbia, Canada 1987-1998 

1-h max O3: 

All yr: 28.02 SD 11.54 

— — — Asthma: 

Age 6-12 yr: 

Range 1.93-105.50 Males: 

Low SES: Lag 1: !35.5% (!52.4, !15.4) 

High SES: Lag 1: !17.8% (!39.6, 11.2) 

Females: Low SES: Lag 1: 32.6% (!7.9, 94.9) 

High SES: Lag 1: !22.5% (!48.9, 14.1) 

Burnett et al. (1999) Toronto, Ontario, Canada 1980-1994 

24-h avg O3: 

All yr: 19.5 IQR 19 

68.4 74.8 0.14-110.8 Asthma: 

All ages: Lag 1-3: 6.5% (3.7, 9.4) 

Lin et al. (2003) Toronto, Ontario, Canada 1981-1993 

1-h max O3: 

All yr: 30.39 SD 17.87 

68.4 74.8 0.14-110.8 Asthma: 

Age 6-12 yr: Males: 

Range 0-141 Lag 0: !7.8% (!22.6, 8.2) 

Females: Lag 0: !26.0% (!39.2, 8.2) 

Burnett et al. (2001) Toronto, Ontario, Canada 1980-1994 

1-h max O3: Summer: 45.2 IQR 25 

77.7 83.7 9-110.8 Total respiratory: 

Age <2 yr: Lag 0: 1.9% (!2.7, 6.8) Lag 0-4: 14.1% (4.9, 24.1) 
Table 8A-4 (cont’d). Effects of Acute O3 Exposure on Total Respiratory and Asthma Hospital A

in the U.S. and Canada 

8A-15

Statistics for 8-h max O3 

Air Quality Data (ppb) a Standardized Percent Exc

Reference, Study 

Location and Period Mean O3 Levels (ppb) 98th % 99th % Range All Year 

Luginaah et al. (2005) 1-h max O3: 78 85 0-106 Total respiratory: Windsor, Ontario, Canada 1995-2000 All yr: 39.3 All ages: 

SD 21.4 Males: Range 1-129 Lag 1: 5.4% (!10.5, 24.2) 

Females: Lag 1: !7.2% (!24.1, 13.5) 

Age 0-14 yr: Males: Lag 1: !7.6% (!33.4, 28.0) Females: Lag 1: 6.7% (!22.7, 47.0) 

Age 15-64 yr: Males: Lag 1: !5.6% (!43.5, 58.0) 

Females: Lag 1: !24.3% (!48.2, 10.5) Age 65+ yr: Males: Lag 1: 13.0% (!10.5, 42.8) 

Females: Lag 1: !7.5% (!29.4, 21.3) 
Table 8A-4 (cont’d). Effects of Acute O3 Exposure on Total Respiratory and Asthma Hospital A

in the U.S. and Canada 

Statistics for 8-h max O3 

Air Quality Data (ppb) a Standardized Percent Exc

Reference, Study 

Location and Period Mean O3 Levels (ppb) 98th % 99th % Range All Year 

8A-16

Schwartz (1996) Spokane, WA Apr-Oct 1988-1990 

1-h max O3: 

Warm season: 41 IQR 12 24-h avg O3: 

— — — — 

Warm season: 29 IQR 9 

Gwynn and Thurston (2001) New York City 1988-1990 

24-h avg O3: 

All yr: 22.1 IQR 14.1 Maximum 80.7 

90.6 106 6-125 Total respiratory: 

All ages: White: Lag 1: 1.1% (-0.4, 2.6) Non-white: 

Lag 1: 4.0% (2.5, 5.6) 

Schwartz et al. (1996) Cleveland, OH Apr-Oct 1988-1990 

1-h max O3: 

Warm season: 56 IQR 28 

91 99 5-120.3 — 

Gwynn et al. (2000) Buffalo, NY 1988-1990 

24-h avg O3: 

All yr: 26.2 IQR 14.8 Maximum 87.6 

92.5 104 4.5-123 Total respiratory: 

All ages: Lag 1: 3.9% (1.8, 6.1) Note: Used Poisson GAM with default convergence criteria 
Table 8A-4 (cont’d). Effects of Acute O3 Exposure on Total Respiratory and Asthma Hospital A

in the U.S. and Canada 

Statistics for 8-h max O3 

Air Quality Data (ppb) a Standardized Percent Exc

Reference, Study 

Location and Period Mean O3 Levels (ppb) 98th % 99th % Range All Year 

Linn et al. (2000) Los Angeles, CA 1992-1995 

24-h avg O3: 

Winter: 14 SD 7 Spring: 32 SD 10 

All yr: 

98.8 

Summer: 

175 

All yr: 

106.9 

Summer: 

180 

All yr: 

4.6-143.2 

Summer: 

13.5-188 

Total respiratory: Age 30+ yr: Lag 0: 1.61% (0.03, 3.22) 

Summer: 33 SD 8 Fall: 15 SD 9

8A-17

a Using O3 data obtained for the study period in the location of the study, 8-h max O3 concentrations were derived and statistics were calculated. The 

 values for the full study period distribution are presented here (unless noted otherwise), along with the range (minimum-maximum) of concentratio

 of the studies vary in length, from several weeks to over 10 yr, the 98th and 99th percentile values were selected for presentation here as a high stud

 that roughly approximates a 4th maximum concentration, depending on the study period length. 

b Percent change in lung function per standard unit ppb O3: 40 ppb for 1-h max O3; 30 ppb for 8-h max O3; 25 ppb for 12-h avg O3; and 20 ppb for 24
Table 8A-5. Effects of Acute O3 Exposure on All-Cause Mortality in the U.S. and Cana

Statistics for 8-h max O3 

Air Quality Data (ppb) a Standardized Percent Exces

Reference, Study Mean O3 Levels 

Location and Period (ppb) 98th % 99th % Range All Year 

8A-18

Vedal et al. (2003) 1-h max O3: 40.5 47.3 1.6-58.7 — All Vancouver, British Lag Columbia, Canada All yr: 27.3 1994-1996 SD 10.2 

Range 3.1!75.1 

Villeneuve et al. (2003) 24-h avg O3: 42.7 47.3 1.1-71.9 Age 65+ yr: Vancouver, British Lag 0: 2.1% (-1.3, 5.3) Columbia, Canada All yr: 13.4 Lag 0-2: 0.4% (-3.2, 4.4) 1986-1999 Range 0.6-38.6 

Fairley (1999; reanalysis 8-h max O3: 67 74 2-105 All ages: Fairley, 2003) Lag 0: 3.0% (!0.3, 6.4) Santa Clara County, CA All yr: 29 1989-1996 SD 15 

Moolgavkar (2003) 24-h avg O3: — — — All ages: All Cook County, IL Lag 0: 0.28% (0.19, 0.36) Lag 1987-1995 All yr: Median 18 

Range 0.2-67 Note: Used Poisson GAM Not

with default convergence wit

Summer: Median 28 criteria. critRange 7-67 

Ito and Thurston (1996) 1-h max O3: 76 85.6 2.7-124 All ages: Cook County, IL Lag 0-1: 3.9% (2.4, 5.8) 1985-1990 All yr: 38.1 

SD 19.9 

Lippmann et al. (2000; 24-h avg O3: 80 85 4.3-101.3 All ages: reanalysis Ito, 2003) Lag 0: 1.84% (!1.73, 5.53) Detroit, MI All yr: 25 Lag 0-3: 0.81% (!3.78, 5.63) 1992-1994 IQR 18-30 
Table 8A-5 (cont’d). Effects of Acute O3 Exposure on All-Cause Mortality in the U.S. and C

Statistics for 8-h max O3 

Air Quality Data (ppb) a Standardized Percent Exces

Reference, Study Mean O3 Levels 

Location and Period (ppb) 98th % 99th % Range All Year 

8A-19

Chock et al. (2000) 1-h max O3: 80 88.9 2.3-92.5 Age 0-74 yr: Age Pittsburgh, PA Lag 0: !1.5% (!5.6, 2.8) Lag 1989-1991 Not reported. 

Age 75+ yr: Age Lag 0: !1.8% (!6.0, 2.6) Lag 

Gamble (1998) 24-h avg O3: 81 86.3 2-98.7 All ages: All Dallas, TX Lag 1-2: 3.7% (0.8, 6.6) Lag 1990-1994 All yr: 22 

Range 0-160 Summer: 30 Range 0-160 

Lippmann et al. (2000; 24-h avg O3: 81.5 88.7 2-123.5 All ages: reanalysis Ito, 2003) Lag 0: 0.60% (!0.62, 1.83) Detroit, MI All yr: 20.9 Lag 0-3: 0.95% (!0.43, 2.35) 1985-1990 IQR 12.0-27.5 

Dockery et al. (1992) 24-h avg O3: — — — All ages: St. Louis, MO Lag 1: 0.6% (!2.4, 3.6) 1985-1986 All yr: 22.5 

SD 18.5 

Dockery et al. (1992) 24-h avg O3: — — — All ages: Eastern Tennessee Lag 1: !1.3% (!7.9, 5.8) 1985-1986 All yr: 23.0 

SD 11.4 
Table 8A-5 (cont’d). Effects of Acute O3 Exposure on All-Cause Mortality in the U.S. and C

Statistics for 8-h max O3 

Air Quality Data (ppb) a Standardized Percent Exces

Reference, Study Mean O3 Levels 

Location and Period (ppb) 98th % 99th % Range All Year 

Lipfert et al. (2000) 7 counties in Philadelphia, PA area 1991-1995 

Schwartz (2005) 14 U.S. cities 1986-1993 

Samet et al. (2000; reanalysis Dominici et al., 2003) 90 U.S. cities (80 U.S. cities with O3 data) 

1987-1994 

1-h max O3: May-Sep: May-Sep: May-Sep: All ages: 

88.8 

All yr: 44.76 SD 25.68 

93.6 2.3-116.6 

Philadelphia: Lag 0-1: 2.49%, p < 0.055 

4 counties in PA:Lag 0-1: 2.52%, p < 0.0557 counties in PA and NJ:Lag 0-1: 2.84%, p < 0.055

1-h max O3: — — — All ages: All 

Lag 0: 0.76% (0.13, 1.40) Lag 

All yr: Median range 35.1 (Chicago, IL) to 60.0 (Provo, UT) 

24-h avg O3: — — — All available data: All 

Lag 

All available data: All ages: Mean range: Lag 0: 0.84% (0.45, 1.24) Approximately 12 Lag 1: 0.38% (0.05, 0.71) (Des Moines, IA) to 36 (San Bernardino, CA) 

8A-20
Table 8A-5 (cont’d). Effects of Acute O3 Exposure on All-Cause Mortality in the U.S. and C

8A-21

Statistics for 8-h max O3 

Reference, Study Mean O3 Levels Air Quality Data (ppb) a Standardized Percent Exces

Location and Period (ppb) 98th % 99th % Range All Year 

Bell et al. (2004) 24-h avg O3: — — — All available data: All 95 U.S. communities Lag 1987-2000 All available data from All ages: Lag 

95 communities: 26 Lag 0: 0.50% (0.24, 0.78) 

Lag 0-6: 1.04% (0.54, 1.55) 

55 communities with all year data: Age < 65 yr: Median range: Lag 0-6: 1.00% (0.20, 1.85) 14.38 (Newark, NJ) to 37.30 (Bakersfield, CA) Age 65-74 yr: 

Lag 0-6: 1.40% (0.56, 2.25) 

40 communities with warm season only data: Age 75+ yr: Median range: Lag 0-6: 1.04% (0.36, 1.75) 20.41 (Portland, OR) to 36.15 (Memphis, TN) 

Klemm et al. (2004) 8-h max O3: — — 6.63-124.41 Age 65+ yr: Atlanta, GA Lag 0-1: 4.2 (!2.4, 11.2) 1998-2000 All yr: 47.03 

SD 24.71 

Moolgavkar (2003) 24-h avg O3: — — — All ages: All Los Angeles County, CA Lag 0: 0.08% (0.01, 0.15) Lag 1987-1995 All yr: Median 24 

Range 0.6-77 Note: Used Poisson GAM Not

with default convergence wit

Summer: Median 36 criteria. critRange 5-77 

Ostro et al. (2000) 1-h max O3: — — — All ages: Coachella Valley, CA !1% (!4, 3) 1989-1998 All yr: 

Note: Used Poisson GAM 

Indio: 62 with default convergence Range 0-180 criteria. Palm Springs: 67 Range 0-190 
Table 8A-5 (cont’d). Effects of Acute O3 Exposure on All-Cause Mortality in the U.S. and C

Statistics for 8-h max O3 

Air Quality Data (ppb) a Standardized Percent Exces

Reference, Study Mean O3 Levels 

Location and Period (ppb) 98th % 99th % Range All Year 

Moolgavkar et al. (1995) 24-h avg O3: — — — — All Philadelphia, PA Lag 1973-1988 Summer: 35.5 

Range 1.3-159.0 

Kinney et al. (1995) 1-h max O3: 115.3 130 5.4-156.1 All ages: Los Angeles County, CA Lag 1: 0.6% (0.0, 1.4) 1985-1990 All yr: 70 

SD 41 

Kinney and Özkaynak 1-h max O3: — — — All ages: (1991) Lag 1: 0.79% (0.33, 1.26) Los Angeles County, CA All yr: 1970-1979 Total oxidants (Ox): 75 

SD 45 

Ostro (1995) 1-h max O3: — — — — All San Bernardino County Lag and Riverside County, Warm season: 140 CA Range 20-370 1980-1986 

8A-22

a Using O3 data obtained for the study period in the location of the study, 8-h max O3 concentrations were derived and statistics were calculate

 99th percentile values for the full study period distribution are presented here (unless noted otherwise), along with the range (minimum-max

 Since the time periods of the studies vary in length, from several weeks to over 10 yr, the 98th and 99th percentile values were selected for p high study period concentration that roughly approximates a 4th maximum concentration, depending on the study period length. 

b Percent change in lung function per standard unit ppb O3: 40 ppb for 1-h max O3; 30 ppb for 8-h max O3; and 20 ppb for 24-h avg O3. c Due to the low mean and relatively small variability in 1-h max O3 concentrations in the study by Vedal et al. (2003), the standardized incre

 may not be appropriate. The effect estimate in the warm season was 11.0% (95% CI: 3.6, 18.6) per mean level (27.3 ppb) increase in 1-h m
Table 8A-6. Toxicological Effects of Acute Ozone Exposure in Animals 

Effect Species 

O3 Concentration and Duration of Exposure a Refer

Inflammation and increased Rat 0.5 ppm × 3 h Bhalla epithelial/endothelial permeability Mouse 0.26 ppm, 8 h/d, 5 d/wk × 1-90 d Kleeb

Dog 0.4 ppm × 6 h Foster 

8A-23

Cardiovascular (  heart rate) Rat 0.1 ppm × 5 h Arito 

(  atrial natiuretic peptide) Rat 0.5 ppm × 8 h Vesely Pulmonary Function (  AHR) Rat 0.05 ppm × 4 h Depuy

Guinea Pig 0.3 ppm × 4 h Segur

(  Vt,  f) Rat b 0.1 ppm × 5 h Arito 

Histological Rat 0.2 ppm, 3-7 d Dorm

Mouse 0.2 ppm, 3-7 d Dorm

Guinea Pig 0.2 ppm, 3-7 d Dorm

Host Defense Rat 0.8 ppm × 3 h Dong 

a Lowest reported ozone concentration from the current literature.b Effects on ventilation only occurred in young rats exposed to 0.1 ppm. No significant changes were seen between young and old ani

0.5 ppm exposures. 
9. ENVIRONMENTAL EFFECTS: OZONE EFFECTS

ON VEGETATION AND ECOSYSTEMS

9.1 INTRODUCTION 

A number of ozone (O3) effect studies were published between 1996 and 2005, and they 

are reviewed in this document in the context of the previous O3 air quality criteria documents 

(AQCDs) (U.S. Environmental Protection Agency, 1978, 1986, 1992, 1996b). The studies 

include multiple plant and tree species, multiple venues, and multiple research approaches from 

empirical data to process models. The research since 1996 continues to support and strengthen 

the conclusions from the previous O3 AQCD, including that: 

(1) the entrance of O3 into the leaf through the stomata is the critical step in O3 effects; 

(2) current ambient concentrations in many areas of the country are sufficient to impair 

growth of numerous common and economically valuable plant and tree species; 

(3) effects can occur with only a few hourly concentrations above 80 ppb; 

(4) a plant’s response to O3 depends upon the cumulative nature of ambient exposure as 

well as the temporal dynamics of concentrations; 

(5) other environmental biotic and abiotic factors (e.g., insects, water, and nutrient 

availability, elevated CO2, and temperature) are also influential to the overall impact 

of O3 on plants and trees. 

Research, to date, has continued to be focused at the species level with very few studies at the 

ecosystem level. Consequently, a high degree of uncertainty remains in our ability to assess the 

impact of O3 on ecosystem services. 

Although chamber exposures still dominate the effects literature, there has been a general 

shift away from chamber studies in favor of more field-based approaches. Field-based 

approaches include surveys of visible injury as well as physiological and growth studies using 

non-chambered free-air CO2 exposure (FACE) systems. The FACE system studies published 

thus far have supported earlier observations of foliar injury and altered growth effects in opentop chambers (OTC) systems. Increased emphasis has also been placed on quantifying aspects 

9-1
of O3 uptake to better link ambient exposure monitoring with plant or tree response. Much of 

this progress has occurred in Europe in developing their O3 air quality management tool, the 

“critical level.” The European research has developed exposure-response functions for several 

European crops and tree seedlings using OTC studies and for use in developing and testing 

models that simulate uptake. Evaluation of the above newly available information has added to 

our knowledge and provides new research directions, but it has not fundamentally altered the 

conclusions of the 1996 O3 AQCD (U.S. Environmental Protection Agency, 1996). 

It is well known that O3 is phytotoxic and that toxicity occurs only if O3 or its reaction 

products reach the target tissues in the plant cell. Recent studies have provided an increased 

understanding of how O3 interacts with the plant at the cellular level. This increased 

understanding of cellular-level O3 effects has been translated into better models, more detailed 

schemata of how O3 alters much of the basic metabolism of plants, and how to construct an 

index that more aptly captures the species, climate, and site factors that alter uptake. These 

results have and will continue to lead to better quantification of exposure and effect. However, 

the translation of these mechanisms into how O3 is involved in altered cell metabolism and 

subsequent reductions in whole-plant productivity and ecosystem-level responses remains to be 

more fully resolved. 

The ensuing sections of this chapter (Sections 9.2 to 9.8) are not intended to provide a 

comprehensive review of the environmental effects of O3, but rather an assessment of key 

information published since the 1996 O3 AQCD. More detailed discussion of the research since 

1996 is provided in Chapter 9 Annex Sections AX9.1 to AX9.7 (in Volume 3 of this document). 

The framework for Chapter 9 follows the environmental effects chapter of the 1996 O3 AQCD. 

First, an overview of various methodologies that have been, and continue to be, central to 

quantifying O3 effects on vegetation is provided in Section 9.2 below (see Section AX9.1 for 

more detailed discussion). The adequacy of each methodology is discussed in the context of 

developing statistically robust data appropriate for assessing and predicting the risk of O3 injury 

to vegetation resources. In Section 9.3, research is reviewed from the molecular to the 

biochemical and physiological levels in impacted plants, offering insight into the mode of action 

of O3 (see also AX9.2). Then, the manner in which plants respond to O3, as influenced by 

9-2
numerous environmental biotic and abiotic factors, is next discussed (see also AX9.3). 

Quantifying these various modifiers is critical to scaling the response of individual plants to the 

community level and across varied landscapes and climates and is needed for regional-tonational assessments of risk. The development of indices of O3 exposure or O3 uptake is 

discussed in the context of their adequacy to realistically describe the ambient concentrationresponse relationships (see also AX9.4). Exposure-response relationships for a large number of 

crop species and cultivars, native vegetation, and tree species are also reviewed, tabulated, and 

compared to form the basis for an assessment of the potential risk from current levels of O3 on 

vegetation resources (see also AX9.5). Available research by which to assess the impact of O3 

on ecosystems is also reviewed, along with data potentially available for estimating the loss of 

various ecosystem services (see also AX9.6). Finally, available research on the economic 

impact of O3 effects on vegetation resources is briefly discussed (see also AX9.7). 

9.2 METHODOLOGIES USED IN VEGETATION RESEARCH 

Methodological advancements since 1996 have not fundamentally altered our 

understanding of O3 effects on plants or ecosystems. Most of the new information confirms 

earlier conclusions and provides additional support for OTC use in assessing plant species and 

developing exposure-response relationships. A more in-depth discussion of this topic can be 

found in Annex Section AX9.1. 

The majority of O3 effects studies are fumigation studies that were conducted in controlled 

chambers, as noted in the 1996 O3 AQCD (U.S. Environmental Protection Agency, 1996). That 

document noted that OTCs represented the best technology for determining statistically robust 

exposure-response models of O3 and crop yield and plant biomass at that time. While OTCs are 

still the best method for conducting multiple, replicated controlled exposures of varying length 

and frequency for developing exposure-response relationships, several new approaches have 

been applied to O3 effects research, most notably free-air exposure systems. FACE systems 

eliminate some concerns about closed or open-top chamber experiments including small plot 

size, altered microclimate within OTCs, and the effect of charcoal filtering on overall air quality 

9-3
within OTCs. The FACE studies have, on the whole, confirmed what was already understood or 

hypothesized about how plants and plant assemblages respond to O3. Some shortcomings of 

using free-air systems in O3 research have also been identified, namely the relatively poor 

control of exposure levels, the presence of “hotspots” and the inability to decrease O3 

concentrations to below ambient levels when ambient concentrations are phytotoxic. 

Nonetheless, the application of FACE systems and other open-air systems to O3-exposure 

research have greatly helped our scaling efforts and are, perhaps, the best current approach for 

studying the response of plant species mixtures to O3 (Nussbaum and Fuhrer, 2000). 

One of the advantages of the application of free-air systems to O3 research is the ability to 

compare response of plants in open-field systems with results from OTCs. In particular, studies 

with quaking aspen (Populus tremuloides L.) performed in OTCand FACE systems and at sites 

along an ambient O3 gradient have showed that O3 foliar symptom expression is generally 

similar across these methodologies, supporting the previously observed level of variation among 

aspen clones in OTC studies (Isebrands et al., 2000, 2001; Karnosky et al., 1999). Recent 

exposure studies using the non-chambered FACE system with soybean cultivars in Illinois 

reported reductions in soybean yield of 15 to 25% (compared to ambient) in two year-long 

studies (Morgan et al., 2006). The results are similar to the reductions reported in multiple 

soybean studies conducted in the 1980s using OTC systems. Multiple-year exposure studies 

employing the FACE system reported foliar injury with reduced volume growth in aspen and 

maple similar to results reported from earlier OTC studies (Karnosky et al., 1999; Isebrands 

et al., 2001). Similar observations between the two exposure systems offer some corroborative 

support for the use of either system. Each has advantages and disadvantages as experimental 

tools and each can be used to effectively investigate O3 effects. Extrapolation of the results from 

chamber studies depends on fully characterizing temperature, light, turbulence, and other 

chamber characteristics during exposures (Nussbaum and Fuhrer, 2000), but study design is 

equally important. Conducting studies with a large number of plant species across regions of the 

country where those species are indigenous is important in delineating regional climatic 

differences in order to reduce the uncertainty associated with extrapolating composited response 

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functions across regions and to identify relative risk to vegetation in relation to given O3 

exposure values (U.S. Environmental Protection Agency, 1996). 

The lack of rural monitors continues to be a major problem in characterizing O3 exposures 

in remote areas and complex landscapes so as to link effects to exposure in natural ecosystems. 

Since the 1996 O3 AQCD, the use of passive samplers has expanded monitoring efforts to 

include remote areas that were previously uncharacterized. This has greatly enhanced our ability 

to link O3 symptomology with elevated O3 exposure in such remote areas. However, passive 

samplers do not capture the temporal dynamics of exposure. Therefore, passive samplers cannot 

substitute for active monitors when attempting to link exposure dynamics to plant response or 

when developing exposure- or dose-response relationships of much value as inputs for the 

standard setting process. To overcome this problem, Krupa et al. (2001, 2003) used models and 

data from a collocated O3 monitor to estimate the underlying frequency distribution of hourly O3 

concentrations from passive samplers. Future development of passive monitor technology and 

data synthesis techniques holds promise, particularly as it is unlikely that extensive O3 

monitoring networks will be established in rural areas in the near future. 

Exclusion methods that employ protective chemicals such as ethylenediurea (EDU) are the 

least disruptive of ambient culture conditions in the field, as noted in the 1996 O3 AQCD. 

However, the level of protection afforded by EDU is site- and species-specific and is subject to 

local meteorologic conditions. In addition, new evidence suggests that EDU does not always 

provide greater protection at higher O3 exposures and that the degree of protection by EDU 

largely depends on environmental conditions. Because of the variability observed in the level of 

protection provided and the fact that mechanisms of protection afforded by EDU and other 

exclusion methods are unknown, caution is needed in applying this approach to the study of O3 

effects in the field. 

Advancements in biomonitoring have been made since the 1996 O3 AQCD, primarily in 

the area of identification and symptom verification of O3-sensitive species (Flagler, 1998; Krupa 

et al., 1998; Innes et al., 2001; Smith et al., 2003). The U.S. Department of Agriculture (USDA) 

Forest Service continues its program to monitor O3 effects in forested ecosystems throughout the 

United States. Currently, 33 states participate in the program, which uses a grid system to 

9-5
identify the location of plants showing foliar injury. Although results cannot be used for 

developing exposure-response relationships or for quantifying responses to O3, they can provide 

an annual assessment and correlative information regarding the extent of O3 injury occurring 

across many regions of the United States. 

9.3 SPECIES RESPONSE AND MODE-OF-ACTION 

Several steps in the process of O3 uptake and toxicity are now better understood than in 

1996. These advancements are important in refining hypotheses on O3 uptake and mode of 

action on plants and in developing a flux-based index for use in quantifying response and, 

ultimately, for potential use in developing a secondary national ambient air quality standard 

(SNAAQS) for O3. The new information available on the mode of action of O3 is, in part, a 

result of improved molecular tools for following rapid changes that occur within the leaf (Ward 

et al., 1991; Pell et al., 1997; Sandermann, 2000). This new information is discussed in greater 

detail in Annex Section AX9.2. 

Clearly, many changes occur within hours or possibly days following O3 exposure 

(Sandermann, 1998). However, other O3 effects take longer to occur (e.g., “carry-over” effects 

seen in the growing season following O3 exposure) and tend to be most obvious only under 

exposure to low O3 concentrations for long periods (Hogsett et al., 1989; Andersen et al., 1997; 

Langebartels et al., 1997). These low-exposure chronic effects have been linked to the 

senescence process or to some physiological response very closely linked to senescence (e.g., 

translocation, reabsorption, storage, and allocation of nutrients and carbon). 

Langebartels et al. (1997) discussed “memory” or carry-over effects within the plant to 

explain sensitivity to frost in the winter following summertime O3 exposure. Others have argued 

that this sensitivity is due to the nutrient status of the tree during the over-wintering phase of its 

life and to chronic (ongoing, less severe levels with fewer peaks at very high levels) exposure to 

ambient O3 inducing (1) mineral nutrient deficiency; (2) alterations of normal metabolism, 

including translocation and allocation of carbohydrates and probably nitrogen; and/or 

(3) disturbance of normal transpiration and diurnal cycling, leading to water stress (Schmieden 

9-6
and Wild, 1995). While generalized low nutrient concentrations within the foliage may not 

occur, localized deficiencies might. This is difficult to observe or to prove without a great deal 

of work on all portions of a tree and without a general hypothesis of what is occurring. 

It is important to note that the dramatic strides made over the last few years in 

understanding the genetic makeup of plants, gene control, and signal transduction and control 

will accelerate in the future and translate into better models of the hypotheses listed above as 

well as into more detailed schemes of how O3 alters basic plant metabolism. Thus, while our 

understanding of how O3 interacts with the plant at the cellular level has dramatically improved 

(Assmann, 2003; Assmann and Wang, 2001; Rao and Davis, 2001), the translation of those 

mechanisms into how O3 is involved with altered cell metabolism and the subsequent reductions 

in whole plant productivity and other physiological facts have not yet been fully achieved. 

As the understanding of wounding responses in plants and more information on genome details 

and varied plant mutants becomes available, the cellular and physiological responses of plants 

to O3 exposures are slowly becoming clearer. However, more studies on a larger variety of 

species are needed before this type of information can be incorporated into indices of response 

and for consideration in developing SNAAQS. 

9.4 MODIFICATION OF FUNCTIONAL AND GROWTH RESPONSES 

It has been known for decades that several factors, both biotic and abiotic, alter plant 

response to O3. However, only a few studies reported since the 1996 O3 AQCD have improved 

our understanding of the role of these interactions in modifying plant O3 response. Quantifying 

how these interactions alter plant O3 response is a critical first step to reducing the uncertainty in 

extrapolating individual plant responses spatially or to higher levels of biological organization, 

e.g., ecosystems. Although the recent studies have not improved our ability to quantify the 

degree to which these factors modify plant O3 response, they have reinforced the conclusions of 

the 1996 O3 AQCD with regard to factors known to alter plant response to O3. This new 

information is discussed in greater detail in Annex Section AX9.3. 

9-7
In the area of biotic interactions, new evidence with regard to insect pests and diseases 

(see Docherty et al., 1997, and Flückiger et al., 2002 for recent reviews) has not altered the 

conclusions and uncertainties noted in the 1996 O3 AQCD. Recent studies have supported the 

earlier conclusion that O3 often increases the likelihood and success of insect attacks, but only 

with respect to chewing insects (e.g., Percy et al., 2002; Kopper and Lindroth, 2003). With the 

economically important group of sucking insects (e.g., aphids), no clear trends have been 

revealed in the latest studies (see reviews by Docherty et al., 1997; Flückiger et al., 2002). 

It seems that some insect problems could increase as a result of greater O3 levels, but predicting 

the likelihood and severity of any particular O3-plant-insect interaction is not possible at 

this time. 

The situation is somewhat clearer with respect to interactions involving facultative 

necrotrophic plant pathogens, with O3 exposure generally contributing to increased disease 

(Flückiger et al., 2002). With obligate biotrophic fungal, bacterial, and nematode diseases, 

however, twice as many reports indicate O3-induced inhibitions as enhancements. This pattern is 

supported by the concept put forth by Dowding (1988) that pathogens that benefit from damage 

to cells are enhanced by pollution stress of their hosts, whereas those pathogens and pests that 

require healthy hosts are depressed by pollution stress. Despite frequent reports that infection by 

obligate biotrophs reduces the severity of O3-induced foliar injury (e.g., Schraudner et al., 1996), 

such infection does not result in true “protection”, as the disease per se causes negative effects 

on the host plant. With obligate biotrophs, the nature of any interaction with O3 is probably 

dictated by the unique, highly specific biochemical relationships between the pathogen and the 

host plant. At this time, therefore, although some diseases may become more widespread or 

severe as a result of exposure to O3, it is still not possible to predict exactly which diseases are 

likely to present the greatest risks to crops and forests. 

Recent studies of interactions between O3 and root symbionts have supported conclusions 

put forth in the 1996 O3 AQCD. Several studies have indicated that the functioning of tree root 

symbioses with mycorrhizae may be adversely affected by O3 (e.g., Kytöviita et al., 2001), 

but there is also evidence that the presence of mycorrhizae may overcome O3-enhanced root 

diseases (Bonello et al., 1993). Also, there is evidence that O3 may encourage the spread of 

9-8
mycorrhizae to the roots of uninfected trees. The role of O3 in altering root symbionts, its 

interactions with soil organisms, and the subsequent feedback effects on plant growth represent 

one of the greatest areas of uncertainty in assessing the influence of O3 on ecosystems and 

ecosystem services (Andersen, 2003). 

The few recent studies of the impact of O3 on intraspecific plant competition confirmed 

that grasses frequently show greater resilience than other types of plants. In grass-legume 

pastures, the leguminous species tend to suffer greater growth inhibition (Johnson et al., 1996; 

Nussbaum et al., 2000). The suppression of Ponderosa pine (Pinus ponderosa Laws.) seedling 

growth by blue wild-rye grass (Elymus glaucus Buckl.) was markedly increased by O3 (Andersen 

et al., 2001). However, predicting the impact of O3 on specific competitive situations, such as 

successional plant communities or crop-weed interactions, is not possible at this time. 

Physical or abiotic factors play a large role in modifying plant response to O3, and new 

information is available that supports the conclusions of the 1996 O3 AQCD. Although some 

recent field studies have indicated that O3 impact significantly increases with increased ambient 

temperature (Ball et al., 2000; Mills et al., 2000), other studies have indicated that temperature 

has little effect (Balls et al., 1996; Fredericksen et al., 1996). Temperature affects the rates of all 

physiological processes based on enzyme catalysis and diffusion; each process and overall 

growth (the integral of all processes) has a distinct optimal temperature range. It is important to 

note that a plant’s response to changes in temperature will depend on whether it is growing near 

its optimum temperature for growth or near its maximum temperature (Rowland-Bamford, 

2000). But temperature is very likely an important variable affecting plant O3 response in the 

presence of the elevated CO2 levels contributing to global climate change. In contrast, evidence 

continues to accumulate that O3 exposure sensitizes plants to low temperature stress (Colls and 

Unsworth, 1992) and, also, that O3 decreases below-ground carbohydrate reserves, which may 

lead to responses in perennial species ranging from rapid demise to impaired growth in 

subsequent seasons (i.e., carry-over effects) (Andersen et al., 1997). 

Light, a component of the plant’s physical environment, is an essential “resource” whose 

energy content drives photosynthesis and CO2 assimilation. It has been suggested that increased 

light intensity may increase the O3 sensitivity of light-tolerant species while decreasing that of 

9-9
shade-tolerant species, but this appears to be an oversimplification with many exceptions. 

As pointed out by Chappelka and Samuelson (1998) and Topa et al. (2001), the interaction 

between O3 sensitivity and light environment is complicated by the developmental stage as well 

as the light environment of individual leaves in the canopy. 

Although the relative humidity of the ambient air has generally been found to increase the 

adverse effects of O3 by increasing stomatal conductance (thereby increasing O3 flux) abundant 

evidence also indicates that the ready availability of soil moisture results in greater O3 sensitivity 

(Mills, 2002). The partial “protection” against the adverse effects of O3 afforded by drought has 

been observed in field experiments and modeled in computer simulations (Broadmeadow and 

Jackson, 2000). There is also compelling evidence that O3 can predispose plants to drought 

stress (Maier-Maercker, 1998). Hence, the response will depend to some extent upon the 

sequence in which the stressors occur; but, even though the nature of the response is largely 

species-specific, successful applications of model simulations have led to larger-scale 

predictions of the consequences of O3-drought interactions. However, regardless of the 

interaction, the net result on short-term growth is negative; although in tree species, other 

responses (such as increased water use efficiency) could benefit long-term survival. 

Mineral nutrients in the soil, other gaseous air pollutants, and agricultural chemicals 

constitute environmental chemical factors. The evidence regarding interactions with specific 

nutrients is still contradictory: some experimental evidence indicates that low general fertility 

increases sensitivity to O3 (Whitfield et al., 1998; Landolt et al., 1997), while others have found 

less sensitivity with decreased fertility (Cardoso-Vilhena and Barnes, 2001). Simulation 

modeling of trees suggests that nutrient deficiency and O3 may act less than additively. 

Interactions of O3 with other air pollutants have received relatively little attention since 

1996 (see Barnes and Wellburn, (1998, and Fangmeier et al., 2002, for recent reviews). The 

situation with SO2 remains inconsistent, but SO2 seems unlikely to pose any additional risk to 

those related to other individual pollutants. With NO and NO2, the situation is complicated by 

their nutritional value as a N source. More information is needed to predict the outcomes of 

different O3-NO-NO2 scenarios. The latest research into O3-acid rain interactions has confirmed 

that, at realistic acidities, significant interactions are unlikely (Laurence et al., 1997; Momen 

9-10
et al., 1997, 1999; Sayre and Fahey, 1999). A continuing lack of information precludes any 

generalizations about interactive effects of O3 with NH3, HF, or heavy metals. More evidence 

was reported for protective effects against O3 afforded by the application of fungicides (Wu and 

Tiedemann, 2002). 

Considerable emphasis during the last decade has been placed on research evaluating 

potential O3 interactions with the components of global climate change: increased atmospheric 

CO2, increased mean global temperatures, and increased surface level UV-B radiation. 

However, it must be noted that most of these studies have tended to regard increased CO2 levels 

and increased mean temperatures as unrelated phenomena. Experiments into the effects of 

doubled CO2 levels at today’s mean ambient temperatures are of questionable value in trying to 

assess the impact of climate change on responses to O3. To date, the limited experimental 

evidence and model simulations suggest that even though an enriched CO2 atmosphere 

(~600 ppm) would more than offset the impact of O3 on responses as varied as wheat (Triticum 

aestivum L.) yield or young Ponderosa pine growth, the concurrent increase in temperature 

would reduce, but probably not eliminate, the net gain (Batts et al., 1997; Van Oijen and Ewart, 

1999; Constable et al., 1996). There is also some recent evidence that O3 and ultraviolet 

radiation of wavelengths 280 to 320 nm (UV-B) interact in their effects on plant injury and 

photosynthesis (Schnitzler et al., 1999), but additional research is needed to fully understand 

how O3 interacts with multiple climate change factors. 

9.5 EFFECTS-BASED AIR QUALITY EXPOSURE INDICES 

Exposure indices are metrics that relate measured plant damage (i.e., reduced growth) to 

monitored ambient O3 concentrations over time to provide a consistent metric for reviewing and 

comparing exposure-response effects obtained from various studies. No new information is 

available since 1996 that alters the basic conclusions put forth in the 1996 O3 AQCD. The 1996 

AQCD (U.S. Environmental Protection Agency, 1996) focused on the research used to develop 

various exposure indices to quantify growth and yield effects in crops, perennials, and trees 

(primarily seedlings) and not foliar injury. The proposed indices included various functional and 

9-11
statistical summaries of monitored hourly O3 concentrations over designated time periods. The 

indices were developed through regression analyses of earlier exposure studies, which was 

accomplished by ordering the measured responses of growth and/or yield of crops and tree 

(seedling) species in response to O3. The indices’ development focused on consideration and 

inclusion of some, but not all, the factors that affect O3 uptake and expression of effects (e.g., 

Lee et al., 1988). 

In the 1996 O3 AQCD (U.S. Environmental Protection Agency, 1996b), it was concluded, 

based on the best available data, that those O3 exposure indices that cumulated differentially 

weighted hourly concentrations were the best candidates for relating exposure to plant growth 

response. Greater weight was given to higher O3 concentrations and daylight hours, and it was 

noted that the timing of peak concentrations and maximum plant conductance was critical in 

determining exposure impact on plants. Various weighting functions were used, including 

threshold-weighted (e.g., SUM06) and continuous sigmoid-weighted (e.g., W126) functions. 

Based on statistical goodness-of-fit tests, these cumulative, concentration-weighted indices could 

not be differentiated one from another. Additional statistical forms for O3 indices have been 

discussed in Lee et al. (1988) and in Chapter 3 of this AQCD. A detailed discussion of effectsbased O3 exposure indices research since 1996 can be found in Annex Section AX9.4. 

The few studies that have been published since the 1996 O3 AQCD continue to support the 

earlier conclusions, including the importance of peak concentrations, and the duration and 

occurrence of O3 exposures in altering plant growth and yield. In addition, a large body of new 

research, mostly out of Europe, addresses the need for an index related to the actual uptake of O3 

by the plant and the flux of O3 from the atmosphere to the O3-affected plant tissues. Despite 

additional research linking estimates of flux with plant response since 1996, information is still 

insufficient to identify a flux-based model that incorporates the necessary complexity across 

space and time to be non-site- or non-species-specific. Based on the current state of knowledge, 

exposure indices that cumulate and differentially weight the higher hourly average O3 

concentrations, but include the mid-level values, still represent the best approach for relating 

vegetation effects to O3 exposure in the United States. 

9-12
The new studies substantiated earlier conclusions on the role of exposure components 

including concentration, duration, and exposure patterns in determining plant growth response 

to O3 (Yun and Laurence, 1999; Oksanen and Holopainen, 2001). Recent studies using different 

exposure patterns have confirmed earlier studies on the role of higher concentrations and 

exposure duration (Nussbaum et al., 1995). A role for higher concentrations is inferred based on 

improved air quality in regions in the western United States (Lefohn and Shadwick, 2000). 

For example, the O3 reductions in the San Bernardino Mountain area since the late 1970s are 

associated with reductions in the higher hourly average O3 concentrations, the number of hours 

of concentrations $0.95 ppm, and the cumulative concentration-weighted exposure index (Lee 

et al., 2003). The mid-range concentrations appeared to be relatively unchanged over the period 

of 1980 to 2000. General forest improvement has been reported following a decrease of O3 

along a decreasing gradient of exposure (Miller and Rechel, 1999; Arbaugh et al., 2003; Tingey 

et al., 2004). These studies suggest that the focus should be on the higher O3 concentrations, 

while including the lower levels, when estimating the effects of O3 precursor emission reduction 

strategies on vegetation. The area has also experienced increasing deposition of N over the same 

time period and, by many indicators, the soil is considered to be N-saturated (Fenn et al., 1996). 

The relative role of N, however, in the measured or simulated tree response in the area has not 

been quantified. 

New studies have demonstrated the potential disconnection of peak events and maximal 

stomatal conductance at xeric to mesic sites in California (Grulke et al., 2002; Panek et al., 2002; 

Panek, 2004). In addition, a few studies have indicated that O3 uptake during nighttime hours is 

greater than previously thought (Grulke et al., 2004; Massman, 2004), and a review of the 

literature suggests a large number of species exhibit some degree of conductance at night 

(Musselman and Minnick, 2000). These studies suggest the need for a reconsideration of 

cumulating exposure 24 h/day and not just during daylight hours in O3 exposure index 

determinations. This lack of coincidence in temporal patterns of conductance and peak ambient 

concentrations introduces uncertainty in assessing the impact of O3. The use of an exposure 

index that does not consider regionally unique climate and site factors modifying stomatal 

conductance may, as a result, under- or overestimate growth effects. The shortcomings of an 

9-13
ambient exposure-based index is especially apparent when assessing the potential impact of O3 

across broad climatic regions of the United States or Europe. Various means to overcome this 

potential problem were addressed in several new studies, and one solution would be to add other 

components to the present statistical summaries of exposure indices (e.g., meteorological) to 

develop flux-based indices. However, the increased biological and meteorological information 

in these indices may make them more regional in their applicability. 

A number of studies have taken a flux-based approach to improve upon the ambient air 

concentration-based (i.e., exposure indices) approach as a means to address the issue of 

assessing risk of O3 across different climatic regions. The European acceptance and use of the 

flux-based critical values is, in part, a recognition of the landscape scaling problems associated 

with ambient exposure-based indices. A great deal of progress has occurred in developing and 

testing stomatal models that may be generally applicable across certain vegetation types 

(Emberson et al., 2000; Pleijel et al., 2000; Danielsson et al., 2003; Grünhage and Jäger, 2003; 

Matyssek et al., 2004). While a flux-based approach is preferred, a cautionary argument has 

been advanced in a few publications, based on the nonlinear relationship between O3 uptake and 

foliar injury (growth was not assessed). The concern is that not all O3 stomatal uptake results in 

a yield reduction, which depends to some degree on the amount of internal detoxification 

occurring with each particular species. Those species having high amounts of detoxification 

potential may, in fact, show little relationship between O3 stomatal uptake and plant response 

(Musselman and Massman, 1999). 

Given the current state of knowledge and the best available data, exposure indices that 

cumulate and differentially weight the higher hourly average concentrations and also include the 

mid-level values continue to offer the most defensible approach for use in developing response 

functions and comparing studies, as well as for defining future indices for vegetation protection. 

A large database exists that has been used for establishing exposure-response relationships; 

however, at this time, such a database does not exist for relating O3 flux to growth response. 

It is anticipated that, as the overlapping relationships of conductance, concentration, and 

defense mechanisms are better defined, the flux-based indices may be able to predict vegetation 

injury and/or damage across varied landscapes and climates with more accuracy than the 

9-14
exposure-response models. However, that is not the case at this time. The translation of these 

indices from research and assessment tools to air quality standards has the additional need to be 

simple, understandable, and adaptive to a manageable monitoring program. 

9.6 OZONE EXPOSURE-PLANT RESPONSE RELATIONSHIPS 

Data published since 1996 continue to support the conclusions of previous O3 AQCDs that 

there is strong evidence that ambient concentrations of O3 cause foliar injury and growth and 

yield damage to numerous common and economically valuable plant and tree species. For 

annual vegetation, the data summarized in Table AX9-16 (see Annex Section AX9.5) show a 

range of growth and yield responses both within species and among species. Nearly all of these 

data were derived from OTC studies, with only two studies using open-air systems in the United 

Kingdom (Ollerenshaw and Lyons, 1999; Ollerenshaw et al., 1999). It continues to be difficult 

to compare studies that report O3 exposure using different indices, such as AOT40, SUM06, 

W126, or 7-h or 12-h mean values. The AOT40, SUM06, and W126 indices are defined as 

follows: 

AOT40: the seasonal sum of the difference between an hourly concentration above 

the threshold value of 40 ppb, minus the threshold value of 40 ppb; 

SUM06: the seasonal sum of hourly concentrations at or above the threshold value 

of 60 ppb; and 

W126: a sigmoid functional weighting of all hourly concentrations for the season. 

When such index comparisons can be made, the results of recent research confirm earlier 

results assessed in the 1996 O3 AQCD (U.S. Environmental Protection Agency, 1996). 

A summary of earlier literature concluded that a 7-h, 3-month mean of 49 ppb O3 corresponding 

to a SUM06 exposure of 26 ppmAh, would cause a 10% loss in 50% of 49 experimental cases 

(Tingey et al., 1991). Recent data summarized in Table AX9-16 support this conclusion and, 

more generally, indicate that ambient O3 exposures can reduce the growth and yield of annual 

species. Some annual species such as soybean [Glycine max (L.) Merr.] are more sensitive, and 

9-15
greater losses in such species may be expected (Table 9-16). Thus, the recent scientific literature 

supports the conclusions of the 1996 O3 AQCD that ambient O3 concentrations are reducing the 

yield of major crops in the United States. 

Much research in Europe has used the cutoff-concentration weighted, cumulative-exposure 

statistic AOT40, and substantial effort has gone into developing “Level-1” critical levels for 

vegetation using this index. Based on regression analysis of 15 OTC studies of spring wheat, 

including one U.S. study and 14 studies from locations ranging from southern Sweden to 

Switzerland, an AOT40 value of 5.7 ppmAh was found to correspond to a 10% yield loss, and a 

value of 2.8 ppmAh corresponded to a 5% yield loss (Fuhrer et al., 1997). Because a 4 to 5% 

decrease could be detected with a confidence level of 99%, 3 ppmAh was selected by the 

European Union as the AOT40 critical level in 1996 (Kärenlampi and Skärby, 1996). 

In addition to reductions in crop yield, O3 may also reduce the quality or nutritive value of 

annual species. Many studies have found O3 effects in various measures of plant organs that 

affect quality, with most of those studies focusing on characteristics important for food or 

fodder. These studies indicate that ambient O3 may have economically important effects on the 

quality of crop and forage species. Previous O3 AQCDs have concluded that visible symptoms 

on marketable portions of crops and ornamental plants can occur with seasonal 7-h mean O3 

exposures of 40 to 100 ppb (U.S. Environmental Protection Agency, 1978; 1986; 1992; 1996). 

The recent scientific literature does not refute this conclusion. 

The use of OTCs may reverse the usual vertical gradient in O3 that occurs within a few 

meters above the ground surface (see Annex Section AX9.1). This reversal suggests that OTC 

studies may, to some degree, overestimate the effects of an O3 concentration as measured several 

meters above the ground. However, such considerations do not invalidate the conclusion of the 

1996 O3 AQCD that ambient O3 exposures are sufficient to reduce the yield of major crops in the 

United States. Recent results from the non-chambered FACE systems, having different vertical 

gradients than chambered systems, report similar results for foliar injury, growth response and 

yield reductions as reported for the chambered OTC systems ( Karnosky et al., 1999; Isebrands 

et al., 2001; Morgan et al., 2006). 

9-16
As found for single-season agricultural crops, yields of multiple-year forage crops are 

reduced at O3 exposures that occur over large areas of the United States. This result is similar to 

that reported in the 1996 O3 AQCD (U.S. Environmental Protection Agency, 1996). When 

species are grown in mixtures, O3 exposure can result in the increased growth of O3-tolerant 

species, which exacerbates the growth decrease of O3-sensitive species. Because of this 

competitive interaction, the total growth of the mixed-species community may not be affected 

by O3 exposure. However, in some cases, mixtures of grasses and clover species have shown 

significant decreases in total biomass growth in response to O3 exposure in studies in the United 

States and in Sweden. In Europe, a provisional AOT40 critical level of 7 ppmAh over 6 months 

has been proposed by the European Union as a value to protect sensitive herbaceous perennial 

plant species from the adverse effects of O3. 

For deciduous tree species, recent evidence from FACE and OTC studies supports results 

observed in previous OTC studies. For example, a series of O3-FACE studies was undertaken in 

Rhinelander, WI (Isebrands et al., 2000, 2001). These studies showed that O3 symptom 

expression was generally similar in OTC and FACE studies and in studies at sites along an 

ambient O3 gradient, supporting the previously observed variation among aspen clones obtained 

using OTCs (Karnosky et al., 1999). As has been observed in previous O3 AQCDs, root growth 

is often found to be the most sensitive biomass response to O3. 

Results of studies since 1996 support the conclusion of the 1996 O3 AQCD (U.S. 

Environmental Protection Agency, 1996) that deciduous trees are generally less sensitive to O3 

than are most annual plants, with the exception of a few very sensitive genera such as sensitive 

clones or genotypes of the Populus genera and sensitive species such as black cherry (Prunus 

serotina Ehrh.). However, the data presented in Table AX9-18 (see Annex Section AX9.5) 

suggest that ambient exposures that occur in the United States can sometimes reduce the growth 

of seedlings of deciduous species. Results from some multiyear studies have shown a pattern of 

increased effects in subsequent years (Hogsett et al., 1989; Anderson et al., 1997; Karlsson et al., 

1995). In some cases, however, growth decreases due to O3 have become less significant or even 

disappeared over time (Karlsson et al., 2002). While some mature trees show greater O3 

9-17
sensitivity than do seedlings in physiological parameters such as net photosynthetic rate, these 

effects may not translate into measurable reductions in biomass growth (Samuelson and Kelly, 

2001). However, because even multiyear experiments do not expose trees to O3 for more than a 

small fraction of their life span and because competition may, in some cases, exacerbate the 

effects of O3 on individual species, determining O3 effects on mature trees remains a significant 

challenge. 

In Europe, a Level I critical level has been set for forest trees based on OTC studies of 

European beech (Fagus sylvatica L.) seedlings and is defined as an AOT40 value of 10 ppmAh 

for daylight hours for a 6-month growing season (Kärenlampi and Skärby, 1996). However, 

other studies show that other species, such as silver birch(Betula pendula Roth.), may be more 

sensitive to O3 than beech (Pääkkönen et al., 1996). 

As found for other tree species, various evergreen tree species and genotypes have widely 

varying O3 sensitivities. Based on OTC studies with seedlings, major evergreen species in the 

United States are generally less sensitive than are most deciduous trees, and slower-growing 

evergreen species are less sensitive than are faster-growing species. There is evidence that plant 

interaction stress, such as competition stress, may increase the O3 sensitivity of trees. As in 

studies of deciduous species, most experiments with evergreen species have only covered a very 

small portion of the life span of a tree and have been conducted with seedlings, so estimating 

effects for mature evergreens is difficult. 

For all types of perennial vegetation, cumulative effects over more than one growing 

season may be important, and studies for only a single season may underestimate effects. 

Mature trees may be more or less sensitive to O3 than are seedlings, depending on the species, 

but information on physiological traits can be used to predict such differences in specific cases. 

In some cases, mature trees may be more sensitive to O3 than seedlings due to differences in gas 

exchange rates or growth rates, to greater cumulative exposures, or to the interaction of O3 with 

other stressors. 

9-18
9.7 EFFECTS OF OZONE EXPOSURE ON NATURAL ECOSYSTEMS 

There is evidence that tropospheric O3 is an important stressor of ecosystems, with 

documented impacts on the biotic condition, ecological processes, and chemical and/or physical 

nature of natural ecosystems (See Table AX9-22; Annex Section AX9.6; Figure 9-1). In turn, 

the effects of O3 on individual plants and processes are scaled up through the ecosystem, 

affecting processes such as energy and material flow, inter- and intraspecies competition, and net 

primary productivity (NPP). Thus, effects on individual keystone species and their associated 

microflora and fauna, effects that have been shown experimentally, may cascade through the 

ecosystem to the landscape level, although this has not yet been demonstrated. By affecting 

water balance, cold hardiness, tolerance to wind and by predisposing plants to insect and disease 

pests, O3 has the potential to impact the occurrence and severity of natural disturbances (e.g., 

fire, erosion). Despite the possible occurrence of such effects, there are essentially no instances 

where ecosystem-level, highly integrated studies have conclusively shown that O3 is indeed 

altering ecosystem structure and/or function. 

Systematic injury surveys demonstrate that foliar injury occurs to O3-sensitive species in 

many regions of the United States (Chappelka et al., 1997; Campbell et al., 2000; Coulston et al., 

2003; Smith et al., 2003) and Europe (Braun et al., 1999). However, the frequent lack of 

correspondence between foliar symptoms and growth effects means that other methods must be 

used to estimate the regional effects of O3 on tree growth rates (e.g., Rebbeck, 1996; Kouterick 

et al., 2000). Investigations of the radial growth of mature trees in combination with data 

from many controlled studies with seedlings and a few studies with mature trees suggest that 

ambient O3 is reducing the growth of mature trees in some locations (Somers et al., 1998). 

Studies using models based on tree physiology and forest stand dynamics suggest that modest 

effects of O3 on growth may accumulate over time and may interact with other stressors 

(Laurence et al., 2001, 2003). For mixed-species stands, such models predict that overall stand 

growth rate is generally not likely to be affected. However, competitive interactions among 

species may change as a result of growth reductions of O3-sensitive species (Weinstein et al., 

2001). These results suggest that O3 exposure over decades may alter the species composition of 

forests in some regions. 

9-19
Figure 9-1. Common anthropogenic stressors and the essential ecological attributes 

they affect. 

Source: Modified from Young and Sanzone (2002). 

Despite the increased understanding of possible ecosystem effects of O3, the database 

demonstrating and quantifying the degree to which O3 is altering natural ecosystems is sparse. 

Much of the speculative O3 impact on ecosystems must be inferred from a number of case 

studies of forest-plot field-based data reporting on a number of different species. One means to 

discuss our current knowledge is by listing the areas in which more information is needed, as 

shown below: 

9-20
Ecosystem processes. Very little is known about the effects of O3 on water, carbon, and 

nutrient cycling, particularly at the stand and community levels. Effects on below-ground 

ecosystem processes in response to O3 exposure alone and in combination with other stressors 

are critical to projections at the watershed and landscape levels. Little is yet known about the 

effects of O3 on structural or functional components of soil food webs or how these impacts 

could affect plant species diversity (Andersen, 2003). 

Biodiversity and genetic diversity. The study of genetic aspects of O3 impacts on natural 

ecosystems has been largely based on correlations, and it remains to be shown conclusively 

whether O3 affects biodiversity or genetic diversity (Davison and Barnes, 1998; Pitelka, 1988; 

Winner et al., 1991). Studies of competitive interactions under elevated O3 levels are needed 

(Laurence and Andersen, 2003). Reexaminations via new sampling of population studies to 

bring a time component into previous studies showing spatial variability in population responses 

to O3 are also needed. These studies could be strengthened by modern molecular methodologies 

to quantify impacts on diversity. 

Natural ecosystem interactions with the atmosphere. Little is known about feedbacks 

between O3 and climate change on the production of volatile organic compounds (VOCs), which, 

in turn, could affect O3 production (Fuentes et al., 2001). At moderate to high O3 exposure 

sites, aberrations in stomatal behavior could significantly affect individual tree water balance 

in O3-sensitive trees, and if the sensitive tree species is dominant, the hydrologic balance at the 

watershed and landscape levels could also be affected. This has not been addressed in any 

model, because O3-exposure effects, if included at all in the modeling effort, have often assumed 

a linear relationship between assimilation and stomatal conductance. Interaction studies with 

other components of global change (i.e., warming, increasing atmospheric CO2, N deposition) or 

with various biotic stressors are needed to better predict complex interactions likely to occur in 

the future (Laurence and Andersen, 2003). Whether O3 will negate the positive effects of an 

elevated CO2 environment on plant carbon and water balances is not yet known; nor is it known 

if these effects will scale up through the ecosystem. How O3 affects the progress of pest 

epidemics and insect outbreaks as concentrations increase is also unclear (Ball et al., 1998). 

Information is needed with regard to the impact of O3 on plant pest and insect reproductive 

9-21
processes and reproductive development under realistic field or forest conditions, as well as 

examination of reproductive effects under interacting pollutants (Black et al., 2000). 

Scaling. The vast majority of O3 tree studies have been conducted with young, immature 

trees and with trees that have not yet formed a closed canopy. Questions remain as to the 

comparability of O3 effects on juvenile versus mature trees, and on trees grown in the open 

versus those in a closed forest canopy, in a competitive environment (Chappelka and Samuelson, 

1998; Kolb and Matyssek, 2001; Samuelson and Kelly, 2001). Merging the effects of O3 across 

spatial scales is also difficult. Scaling responses of a single or a few plants to effects on 

communities and ecosystems are complicated matters that will require a combination of 

manipulative experiments with model ecosystems; community and ecosystem studies along 

natural O3 gradients; and extensive modeling efforts to project landscape-level, regional, national 

and international impacts of O3. Linking these various studies via impacts on common research 

quantification across various scales using measures of such factors as leaf area index or spectral 

reflective data, which could eventually be remotely sensed (Kraft et al., 1996; Panek et al., 

2003), would provide powerful new tools for ecologists. 

Identifying endpoints. In general, methodologies to determine the important values of 

services and benefits derived from natural ecosystems are lacking. Identifying and quantifying 

factors that could be used in comprehensive risk assessment for O3 effects on natural ecosystems 

would increase societal awareness of the importance of protecting ecosystems (Heck et al., 

1998). 

9.8 ECONOMICS 

Substantial progress has been made over the past two decades in our understanding of the 

effects of O3 and other oxidants on vegetation, particularly for agriculturally important plant 

species (see Annex Section AX9.7 for a more detailed discussion). The physical and economic 

effects on agriculture are well documented and provide useful information for consideration in 

establishing air quality standards for crops (e.g., Spash, 1997). 

9-22
Since the completion of the National Crop Loss Assessment Newwork (NCLAN) program 

in the late 1980s, the number of economic assessments of air pollution studies focusing on 

terrestrial ecosystems in general, and agriculture in particular, has declined. For example, 

33 economic studies of O3 and other air pollutant effects on U.S. crops were published in 

peer-reviewed journal outlets from 1980 to 1990 (Spash, 1997). However, in preparing this 

section of the current O3 AQCD, only four peer-reviewed economic assessments addressing 

vegetation in the United States were found for the decade of 1991 to 2000. In addition, one 

peer-reviewed article (Kuik et al., 2000) was found dealing with agriculture in the Netherlands. 

Recent interest in global climate change and the potential effects of global warming on O3 and 

other photochemical oxidants has renewed interest in the effects of air pollution on both 

managed and unmanaged terrestrial ecosystems (Adams et al., 1998). In addition, concern is 

growing regarding the effects of air pollutants on natural ecosystems and on the services they 

provide (Daily, 1997). Unfortunately, this interest has not yet translated into additional 

peer-reviewed publications addressing O3 or other air pollutants’ effects on ecosystems. 

A study by Murphy et al. (1999) of the economic effects of tropospheric O3 on U.S. 

agriculture is of note here, because it confirms the general magnitude of economic effects 

reported by the two key studies performed a decade earlier (Adams, 1985, 1986). Specifically, 

Murphy et al. (1999) evaluated benefits to eight major crops associated with several scenarios 

concerning the reduction or elimination of O3 precursor emissions from motor vehicles in the 

United States. Their analysis reported a $2.8 billion to $5.8 billion (1990 dollars) benefit from 

complete elimination of O3 exposures from all sources, i.e., ambient O3 reduced to a background 

level assumed to be 0.025 to 0.027 ppm. While the analytical framework is similar to Adams 

et al. (1986) in the use of NCLAN-based yield response functions and a mathematical 

programming-based economic optimization model, the study is novel in its focus on the role of 

motor vehicle emissions of VOCs and nitrogen oxides (NOx) in total anthropogenic O3 levels. 

The study is also notable in its careful attention to federal farm program effects, particularly the 

deficiency payment component. 

9-23
A number of recent studies of air pollutant effects on tree species have appeared in the 

literature. Some studies have reported changes in total biomass and focused on European 

species (Kurczynska et al., 1997). Other studies have assessed changes in composition of forest 

species (biodiversity) or forest health due to exposure to air pollutants (Bringmark and 

Bringmark, 1995; McLaughlin and Percy, 1999; Vacek et al., 1999). As noted previously, 

changes in forest biomass and composition are more difficult to value than marketable products. 

However, measures of forest composition or health have implications for an area of increasing 

policy concern, that being the effect of air pollutants and other environmental stressors on 

unmanaged (natural) ecosystems and the services they provide (Goulder and Kennedy, 1997; 

Pimentel et al., 1997). Considerable discussion has occurred among ecologists and economists 

as to the appropriate means for valuing these services (Carpenter and Dixon, 1985; Anderson, 

1990; Common and Perrings, 1992). A number of conceptual articles have been published on 

this issue in both economic and ecological journals (Bergstrom, 1990; Suter, II, 1990; Castle, 

1993; Pearce, 1993). 

Effects on forests and natural ecosystems remain problematic because of limitations in 

biological response data and economic methods. The problem is even more acute for valuing 

natural ecosystem service flows. The current limitations surrounding forests and natural 

ecosystems present a rich research agenda. Areas of greatest potential value in terms of regional 

policymaking need to be prioritized. Such priority setting can be assisted by sensitivity analyses 

with existing economic models. By measuring the changes in economic effects arising from 

changes in key parameters, it is possible to identify those research data gaps most likely to affect 

economic values. 

9-24
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10. THE ROLE OF TROPOSPHERIC OZONE

IN UVB-RELATED HUMAN HEALTH OUTCOMES

AND IN CLIMATE CHANGE

10.1 INTRODUCTION 

In addition to directly affecting human health, vegetation and ecosystems, tropospheric 

ozone (O3) influences the ground-level flux of solar ultraviolet (UV) radiation and alters the 

Earth’s radiative balance by functioning as a greenhouse gas, therefore contributing to climate 

change. This chapter addresses these effects. 

10.2 THE ROLE OF TROPOSPHERIC OZONE IN DETERMINING 

GROUND-LEVEL UV-B FLUX 

Atmospheric O3 plays a crucial role in reducing the exposure of living organisms to solar 

UV radiation. The stratospheric O3 layer is responsible for nearly all of this shielding effect, as 

90% or more of the total atmospheric burden of O3 is located there. Specific quantification of 

the role that tropospheric O3 plays in screening the earth’s surface from harmful UV radiation 

has not been accomplished, to date. 

Reasonable estimation of the importance of ground-level O3 in US urban and suburban 

areas, i.e., the fraction of tropospheric O3 subject to regulation by the US NAAQS, in altering 

human UVB exposure and the incidence of UVB-induced human diseases requires an explicit 

accounting for key factors that influence (a) ground-level flux in US urban and suburban areas, 

(b) the probability of UV-B exposure experienced by sensitive populations, and (c) links

between exposure and the incidence rates of UVB-induced diseases. This section summarizes, 

in this order, available information on such factors. 

10.2.1 Factors Governing Ultraviolet Radiation Flux at the Earth’s Surface 

Given its role in protecting life from harmful UV radiation, photochemical processes that 

alter the concentration of stratospheric O3 are of particular concern to the global community. 

Scientific understanding of the significant losses of stratospheric O3 due to chemistry with the 

10-1
degradation products of the long-lived anthropogenic chlorinated- and fluorinated-hydrocarbons 

evolved during the 1970s and early 1980s. This realization led to the international treaty for the 

protection of stratospheric O3, i.e., the 1987 Montreal Protocol on Substances that Deplete the 

Ozone Layer. The scientific community continues to track stratospheric O3 levels and to 

document any changes in ground-level UV-B flux due to anthropogenic depletion. An outcome 

of these efforts is a limited body of literature that describes the effects of tropospheric pollutants, 

particulate matter (PM), clouds and O3 on ground-level UV radiation flux. 

The Montreal Protocol requires routine review of the latest scientific information available 

on the status of the O3 layer and of UV radiation levels at the Earth’s surface. The World 

Meteorology Organization (WMO) and U.N. Environmental Program (UNEP) are responsible 

for assessing the state of the science regarding the O3 layer and for reporting on trends in surface 

UV radiation levels. The latest WMO/UNEP assessment was published in 2002 (WMO, 2002) 

and includes an evaluation of the role of the troposphere in determining UV flux at the earth’s 

surface. 

The mixture of gases, clouds, and particles that comprise the troposphere scatter and/or 

absorb incident solar radiation (see Figure 10-1). In general, these effects are greater in the 

troposphere than in the stratosphere, due to the higher atmospheric pressures, particle and cloud 

densities present there. Solar flux intensity has a temporal dependence, while radiative 

scattering and absorption have strong wavelength, pathlength, and/or particle concentration 

dependencies. These combine to create nonlinear effects on UV flux at the earth’s surface. 

Thus, careful quantification of atmospheric absorbers and scatterers, along with a well-resolved 

description of the physics of these interactions, is necessary for predicting the impact of groundlevel O3 on UV-B flux. The following sections (10.2.1.1–10.2.1.4) summarize the relevant 

physics that govern solar flux and the nature of radiative interactions in the atmosphere. The 

sources for this information are a selection of atmospheric physics and chemistry texts, the 

WMO (2002) report, and other peer-reviewed literature on the role of the troposphere in defining 

UV surface flux. 

10.2.1.1 UV Radiation: Wavelengths and Energies 

The energy possessed by a photon is inversely proportional to its wavelength. For 

example, gamma rays, having wavelengths ~0.1 nm, are especially damaging high-energy 

10-2
Figure 10-1. Complexity of factors that determine human exposure to UV radiation. 

In addition to the geophysical/atmospheric factors (e.g., stratospheric and 

tropospheric O3, clouds, aerosols, and Rayleigh scattering) that affect the 

solar flux of UV radiation at surface level, there are human physical, 

behavioral and demographic factors that influence human exposure to 

UV radiation. 

photons (>106

 kJ einstein!1) emitted during radioactive decay and by stellar activity. 

Radiowaves, having wavelengths ~108

 nm, are very low in energy (~10-3–10-8 kJ einstein!1) 

and function as carriers for broadcast communications (Finlayson-Pitts and Pitts, 1986). 

The wavelengths ranging between 50 and 400 nm in length are denoted “ultraviolet,” 

and are subdivided into classes, where UV-C corresponds to wavelengths <280 nm, UV-B refers 

to the range 280–320 nm, and UV-A to 320–400 nm. UV-C is almost entirely blocked by the 

Earth’s upper atmosphere, where it participates in photoionization and photodissociation 

processes. Solar UV-B radiation is absorbed or scattered in part within the atmosphere. UV-A 

radiation can be scattered but is not absorbed to any meaningful degree by atmospheric gases. 

Both UV-B and UV-A photons contain sufficient energy to break (photolyze) chemical 

bonds in biomolecules and are associated with human health and ecosystem damage. However, 

10-3
because UV-B is more energetic, it is capable of producing substantially more biological damage 

than UV-A. 

10.2.1.2 Temporal Variations in Solar Flux at the Earth’s Surface 

The magnitude of the solar radiation flux entering the atmosphere depends upon long-term 

solar activity, sunspot cycle (11 years), solar rotation (27 days), and the position of the Earth in 

its orbit around the sun. At any given point on the Earth’s surface, solar flux is dependent upon 

solar zenith angle, a quantity that varies with time of day, season, and latitude. 

Solar Rotation and the Sunspot Cycle 

A variety of changes in solar irradiance can be found in historical data, from 1700 to the 

present. Solanki and Fligge (2000) concluded that solar irradiation changes, on time-scales of 

days to centuries, can be attributed to variations in solar magnetic features. The maximum level 

of radiation (solar-max) differs from the minimum (solar-min) by as much as 10% for 

wavelengths near 160 nm. This peak-to-trough difference declines to ~1% for 300 nm (Salby, 

1996). 

On the decadal time scale, solar rotation and sunspot activity have the largest effects on 

overall solar irradiance. Since the last Maunder minimum in 1700, and consistent with 

increasing sunspot activity, solar irradiance has increased by ~3.0% for wavelengths in the 

UV-C range and by ~1.3% for wavelengths in the UV-B and UV-A ranges. Including visible 

wavelengths, Solanki and Fligge (2000) estimated that the overall increase in solar irradiance 

was ~0.3%. Rozema et al. (2001) pointed out that any increase in wavelengths <300 nm (UV-C) 

would initiate additional O3 formation in the stratosphere. This suggests that any increase in 

UV-B and/or UV-A solar flux could, therefore, be offset by a more absorptive stratosphere. 

The Position of the Earth with Respect to the Sun 

The combined effects of the Earth’s obliquity (the angle of the Earth’s axis of rotation with 

respect to the plane of its orbit around the sun) and its precession (the rotation of the Earth’s axis 

with respect to a perpendicular line through the plane of its solar orbit) yield variations of up to 

30% in total summertime solar flux at the top of the earth’s atmosphere, depending on latitude 

(Hartmann, 1994). 

10-4
Zenith Angle: Latitude, Season, and Time of Day 

The sun’s relative elevation from a point on the earth’s surface is measured with respect to 

the vertical and is known as its solar zenith angle (SZA). This angle varies hourly, seasonally, 

and with latitude. Diurnal and seasonal changes in SZA result in the largest changes in the 

magnitude of solar radiation flux at the surface, with larger zenith angles corresponding to lower 

solar flux. The largest natural fluxes occur in the tropical regions, where solar noon occurs at an 

SZA at or near 0°. Seasonal variation in solar flux ranges from small changes at the equator to 

very large changes at high latitudes. Diurnal variations in solar flux, from sunrise to sunset, 

show added wavelength dependence as a function of SZA. This wavelength dependence is a 

function of the concentrations of radiation scattering and absorbing gases and particles in the 

atmospheric pathlength traversed by the photon. These processes will be discussed further 

below. 

10.2.1.3 Atmospheric Radiative Interactions with Solar Ultraviolet Radiation 

When solar radiation enters the earth’s atmosphere, it can be either scattered or absorbed 

by the gases and particles it encounters. The main mechanisms of atmospheric scattering are 

Rayleigh, in the case of scattering by gas molecules, and Mie, in the case of scattering by aerosol 

particles. The intensity of the Rayleigh effect, i.e., the extent to which the gas molecule is 

capable of perturbing the trajectory of the incident photon, depends on the size of the gas 

molecule in relation to the photon, and is maximized when the photon is comparable in size or 

smaller than the gas molecule. Light is scattered symmetrically with respect to the gas molecule, 

with similar forward and backward intensities. This effect explains the color of the clear sky, 

as blue photons are comparable in size to the dominant gases in the atmosphere (N2 and O2). 

With Mie scattering, the result of light encountering larger particles, photons are scattered with a 

strong forward tendency. This tendency explains why dense water clouds, i.e., cumulus clouds, 

appear brilliantly white. 

The lower atmospheric pressures in the stratosphere mean fewer gas molecules are present 

that can absorb or scatter radiation. Stratospheric clouds and aerosols are also thinner and more 

dispersed than those in the troposphere. In the language of the radiative transfer literature, these 

conditions make the stratosphere a “single scattering” regime for UV radiation. The 

troposphere, due to its high gas and particle concentrations is referred to as a “multiple 

10-5
scattering” regime. In practical terms, UV radiation traverses the stratosphere with a 

substantially lower probability of encountering a gas molecule or cloud or an aerosol particle 

than it would in the troposphere. 

The multiple scattering of solar photons in the troposphere accounts for the 

“disproportionate” role that tropospheric O3 is said to play in absorbing UV radiation versus 

stratospheric O3 on a molecular per molecule basis (Brühl and Crutzen, 1989; Balis et al., 2002; 

Zerefos et al., 2002). The probability that an individual photon will encounter an O3 molecule is 

increased when its effective pathlength through the atmosphere is increased with multiple 

scattering. This leads to an increase in the probability of absorption by O3. This effect is well 

understood in principle, but its importance has not been quantified in relation to stratospheric O3 

or with respect to other tropospheric absorbers. 

The importance of atmospheric pressure and particle concentrations is evident in the 

altitude dependence of solar flux. Solar flux has been observed to increase with altitude above 

sea level, consistent with an inverse relationship between cloud levels, air pollution 

concentrations, and intensity of Rayleigh scattering with altitude. A number of measurements of 

UV radiation have been taken at various altitudes and are reviewed by Xenopoulos and Schindler 

(2001). Increases in flux as a function of altitude are given as percent irradiance enhancement 

per 1000 m relative to sea level. The effect can range from 9 to 24% /1000 m, depending upon 

the altitude at which the measurement was taken (Xenopoulos and Schindler, 2001). 

Similarly, this effect is seen as a function of SZA. For example, at very small zenith 

angles, near solar noon, photons have a shorter atmospheric pathlength to the surface and a 

smaller probability of encountering gas molecules, clouds, or particles. Conversely, photons that 

enter the atmosphere at larger zenith angles, i.e., increasingly tangential to the surface, are more 

likely to be scattered or absorbed. 

Radiative Interactions in the Stratosphere: Absorption by Ozone 

As noted earlier, the stratosphere contains 90% or more of the total column O3. Ozone 

interacts with UV radiation by scattering the photon, or absorbing it. Photoabsorption by O3 

occurs with very high efficiency (Note: for a discussion of the metrics used and relative 

efficiencies of photoabsorption by gas-phase molecules, the reader is referred to text books such 

as Finlayson-Pitts and Pitts [1986]). After electronically-excited O3 (O3*) is formed, it can 

10-6
dissociate into ground-state oxygen, O2, and an electronically excited oxygen radical, O(1

D) 

(see Reaction 10-1). At low pressures and with irradiation at 300 nm (UV-B range), O(1

D) is 

produced with a quantum efficiency near unity. 

(10-1) 

(10-2) 

However, in the presence of other atmospheric gases, intermolecular collisions may also 

disperse the excess electronic energy of the O3* molecule by transferring it to other molecules as 

vibrational, rotational, and translational energies, warming the atmosphere, and reducing the 

production of O(1

D) (see Reaction 10-2). The efficiency of the energy transfer process is a 

function of atmospheric pressure and the quantum mechanical properties of the colliding 

molecules. A deeper discussion of molecular energy transfer processes can be found in the text 

by Levine and Bernstein (1987). 

The Dobson Unit (DU) is conventionally used for discussing stratospheric O3 

concentrations. One DU = 2.687 × 1016 molecules of O3/cm2. Alternatively, a DU corresponds 

to the column height in hundredths of a millimeter of O3 at standard temperature and pressure 

(273 K and 1 atmosphere), integrated along the total height of the atmospheric column 

(Wayne, 2000). 

As previously noted, the total O3 column effectively prevents any UV-C from reaching the 

surface and reduces the penetration of UV-B to the surface, but it does little to attenuate the 

intensity of UV-A except at the shorter wavelengths close to the cutoff for UV-B. Figure 10-2 

compares the solar flux above the atmosphere with ground-level flux. Cutchis (1974) calculated 

that with overhead sun, a 10% decrease in the O3 column would lead to 20, 250, and 500% 

increases in flux at 305, 290, and 287 nm, respectively. These estimates have been supported by 

ground observations in Toronto, ON (49° N; Kerr and McElroy, 1993). Rapid changes of this 

magnitude appear to happen naturally. As seen in data collected by the Total Ozone Mapping 

Satellite (TOMS) (Figure 10-3), the total O3 column undergoes wide natural variation on short 

timescales (Cockell, 2001). 

10-7
Figure 10-2. Comparison of solar flux above the atmosphere with flux at the Earth’s 

surface. The dotted line represents extraterrestrial solar flux measured by 

the satellite UARS SOLSTICE instrument (dotted line). The dashed line 

represents calculated atmospheric transmittance and the solid line is the 

calculated absolute flux of UV radiation for a solar zenith angle of 50°, total 

column O3 of 275 DU, and a surface reflectivity of 8%. The fine structure on 

the surface flux trace results from Fraunhofer lines (absorption specific 

wavelengths within the solar atmosphere). 

Source: Krotkov et al. (1998). 

The WMO (2002) assessment reported that global average total column O3 had declined by 

3% from pre-1980 levels, due to the presence of anthropogenic O3-depleting substances in the 

atmosphere. Ozone depletion has a strong latitude and seasonal dependence. The seasonality of 

total O3 changes differs between the Northern and Southern Hemispheres. In the northern 

midlatitudes, total column O3 declined by ~4% during the winter/spring seasons and by 

approximately half that amount in the summer/fall of the 1997–2001 time period, relative to pre1980 total column O3 levels. In southern midlatitudes, total column O3 declined ~6% during all 

seasons, possibly tied to the enhanced photochemical losses associated with the meteorological 

dynamics peculiar to the antarctic region. 

10-8
Figure 10-3. Ozone column abundances from the years 1990 to 1992 for 0, 40, and 80° N 

as well as 80° S. The data for 80° S are incomplete, but the graph shows the 

effects of the Antarctic O3 hole on total column abundances at this latitude. 

The data for the Northern Hemisphere illustrate the natural variations in the 

O3 column over time. The data are taken from the TOMS (Total Ozone 

Monitoring Satellite) data set (1979 to 1993). 

Source: Cockell (2001). 

Radiative Interactions in the Stratosphere: Scattering by Clouds and Particles 

During periods of extreme cold in the stratosphere above the Earth’s poles, nacreous or 

polar stratospheric clouds (PSCs) form due to the condensation of sulfuric and nitric acids or 

water vapor. Like tropospheric water clouds, PSCs have the capacity to scatter incident solar 

radiation. Due to their infrequent appearance, location, and limited optical density, they are not, 

however, an important factor in determining ground-level UV flux for human health impacts 

assessment. 

Stratospheric aerosols such as those produced by explosive volcanic eruption, i.e., 

Mt. Pinatubo, have the potential for playing a greater role in altering ground-level UV flux. 

In the case of Mt. Pinatubo, a large quantity of sulfur dioxide was injected into the stratosphere, 

which reacted to form a layer of sulfate aerosols. These aerosols increased the Earth’s albedo for 

10-9
a period of years, reducing solar flux at the Earth’s surface. Volcanic eruptions that lead to the 

injection of aerosols containing light-absorbing-material, such as crustal material (dust), can also 

reduce surface flux (Herman et al., 1997). However, these events are also too infrequent to be of 

importance in this discussion. 

Radiative Interactions in the Troposphere: Gases, Particles, Clouds, and the Earth’s Surface 

The components of the troposphere that are of most importance in relation to the prediction 

of ground-level UV radiation flux are clouds, PM, and UV-absorbing trace gases. The 

UV-shielding efficiencies of these components are highly dependent on density (concentration), 

particle size, and the altitude at which they are present in the troposphere. All of these properties 

are sensitive to meteorology, which introduces an element of temporal dependency. The 

following section discusses some of the issues specific to each component. 

Gases 

UV-absorbing gases, including O3, NOx, VOCs, and SO2, are vented into the upper 

troposphere, injected downwards from the stratosphere (O3) or form in situ, i.e., O3, VOC 

oxidation products (see Chapter 3). Ozone concentrations decrease with increasing altitude from 

the surface up to, roughly, the mid-troposphere, then increase up into the stratosphere. 

Figure 10-4 shows a series of O3 vertical profiles for 4 sites within the continental U.S., i.e., 

plots of O3 concentrations as a function of atmospheric pressure (correlating to altitude). The 

mean values of O3 in the free troposphere reported in the literature range from ~50 to ~80 ppbv, 

with higher values occurring at the tropopause. For example, a series of ozonesonde soundings 

over France from 1976 to 1995 showed an O3 increase from 48.9 ppbv in the 2.5 to 3.5 km layer 

to 56.5 ppbv in the 6.5 to 7.5 km layer (Ancellet and Beekmann, 1997). 

Within the planetary boundary layer, photochemistry produces a diurnal rise and fall in O3 

and PM concentrations, especially in polluted urban settings. Temperature inversions that often 

occur in these settings prevent the upward mixing and dilution of ground-level O3, also trapping 

primary and secondary PM within the boundary layer. These effects are described in Chapters 2 

and 3 of this document. Such conditions substantially increase the UV absorptive capacity of the 

atmosphere, at the surface. 

10-10
Figure 10-4. Monthly averaged vertical O3 profiles (partial pressure in mPa) as a function 

of atmospheric pressure (in mBar) for Trinidad Head, CA (solid line); 

Boulder, CO (dot-dashed line); Huntsville, AL (dotted line); and Wallops 

Island, VA (dashed line). The number of launches at each site for each 

month are indicated on the charts. 

Source: Newchurch et al. (2003). 

10-11 
Ultraviolet absorption by gases becomes significant under aerosol- and cloud-free 

conditions. Figure 10-5 shows a calculation by Krotkov et al. (1998) of the sensitivity, as a 

function of wavelength, of ground-level UV flux to a 1-DU decrease in total column O3 under 

cloud- and aerosol-free conditions. (Note: 1 DU = 2.687 × 1016 molecules of O3 per cm2

 or if 

calculated for the entire troposphere, 1 DU = 10.9 Tg(O3) with 1 ppb of tropospheric O3 = 

0.65 DU). 

A study in Chicago in 1991 and 1992, during which ambient O3, broadband UV irradiance, 

and total sunlight were monitored (Frederick et al., 1993), found a significant negative 

correlation between the UV irradiance and ambient O3 when the atmosphere was relatively free 

of clouds and haze. Although Frederick et al. (1993) estimated that a 10 ppbv reduction in O3 

was associated with a 1.3% increase in erythemally-weighted UV-B, they cautioned that this 

value had a comparatively large uncertainty (~1.2%, or nearly 100% of the estimated increase). 

Matthijsen et al. (2000) noted that scattering effects differ for UV versus total solar irradiance: 

UV radiation, even under clear sky conditions is mostly diffuse, whereas total solar is largely 

direct. This is consistent with Rayleigh scattering of UV wavelengths by N2 and O2. 

Clouds 

Clouds have the largest observed influence on solar irradiance–scattering as much as 100% 

of incoming direct beam radiation versus the 10–20% scattering that occurs on clear days due to 

Rayleigh and PM. The long term average for the U.S., based on measurements taken at various 

locations, show monthly average reductions in UV levels ranging from 10% to 50%, due to 

cloud cover (Madronich et al., 1995). An overall reduction in normalized global irradiance has 

been attributed, in part, to the increase in the extent (8%) and optical density of cirrus clouds 

observed since the middle of the last century (Trepte and Winkler, 2004). 

The depth, particle size distribution, and coverage determine, in large part, the amount and 

wavelengths of radiation that clouds will scatter. In general, cloud effects are weaker for the UV 

wavelengths than for total solar radiation. Alados-Arboledas et al. (2003) observed that the 

cloud effect under overcast conditions for total solar radiation is 33% greater than for 

erythemally-weighted UV radiation. 

Geometry is also an especially important factor: scattered or broken clouds can either 

slightly reduce or enhance irradiance due to scattering between clouds (WMO, 2002). 

10-12
Figure 10-5. The sensitivity of ground-level UV flux to a 1 DU change in total column O3, 

under clear sky conditions, as a function of solar zenith angle (SZA). 

Source: Krotkov et al. (1998). 

Palancar and Toselli (2004) estimated that broken clouds can increase noontime radiation by up 

to 34% when compared to calculated clear sky values. Mims and Fredrick (1994) reported 

observations of UV-B radiation enhancements of up to 20% over maximum noontime values due 

to scattering from cumulus clouds. Overcast conditions have been estimated, for SZAs 

averaging 50°, to reduce UV-B surface radiation to 29% of clear sky irradiances (Schafer et al., 

1996). Schafer et al. (1996) also observed, as a percentage of clear sky values, 61% for 80–90% 

coverage, 75% for 60–70% coverage, 79% for 40–50% coverage, with frequent incidences of 

enhancement above clear sky levels due to reflections during partial cloud cover. 

The altitude at which clouds are located also substantially alters their effectiveness in 

scattering radiation back into space. High clouds have been observed to be more effective, by a 

factor of 5, at scattering 300 nm wavelength radiation (SZA = 0°) than low clouds (Wen and 

Frederick, 1995). At SZA = 60°, 305 nm wavelength radiation scatters from high clouds 

10-13
(8–10 km) at twice the efficiency of low clouds (1–2 km). The authors attribute this effect to 

the reduction in the atmospheric pathlength that UV photons must travel, and therefore a 

reduction in the odds of scattering due to other particle or gas interactions, before encountering 

the cloud. 

Quantifying the effect of clouds on surface UV flux, therefore, requires detailed 

information on cloud depth, particle size distribution, geometry, altitude, and the position of the 

sun relative to the cloud as a function of time. 

Particulate Matter 

On a zonally averaged basis, PM does not contribute significantly to lower tropospheric 

absorption of UV radiation. However, in urban areas or other areas subject to high smog levels, 

such as areas adjacent to significant biomass combustion, PM may be the most important 

determinant of ground-level erythemal UV flux, excepting clouds (U.S. Environmental 

Protection Agency, 2004; WMO, 2002). Model-to-measurement comparisons of ground-level 

flux for Cordoba City (Argentina), the Aegean Sea (Greece) and Toronto (Canada), have shown 

20%, 20% and 5–10% reductions, respectively (McKenzie et al., 2003). Liu et al. (1991) 

estimated similar potential reductions (20%) in UV-B radiation in the presence of large local PM 

concentrations. 

Barnard et al. (2003) measured UV-B penetration to the surface at 3 sites: Rubidoux 

(California); the University of California, Riverside; and Mt. Gibbes in the North Carolina Blue 

Ridge Mountains. They found that UV-B is primarily sensitive, under cloud-free conditions, 

to O3 values measured by satellite, which do not account for ground level O3, and concentrations 

of PM and specifically PM containing black carbon. They found that ground level O3 

concentrations did not meet the 0.05 significance level in their model for urban environments. 

Barnard et al. (2003) further suggested that increases over the past 20 to 30 years in combustionassociated PM and black carbon may account for the inability to detect a surface trend in UV-B 

radiation caused by a known decrease in stratospheric O3 over the Northern Hemisphere. 

Surface Albedo 

The Earth’s surface, by absorbing or reflecting radiation, directly influences UV flux in the 

lower troposphere. The extent to which the surface reflects incident radiation back into the 

10-14
atmosphere is known as its “albedo.” Ultraviolet flux is directly proportional to surface albedo 

(Wendisch and Mayer, 2003). 

Surface albedo is a very strongly wavelength-dependent property. For example, fresh and 

wet snow reflect 60 to 90% of incident violet light, while soil and grass surfaces reflect <5% 

(Xenopoulos and Schindler, 2001). In their in situ measurement and modeling study of the 

vertical distribution of solar irradiance, Wendisch and Mayer (2003) found that surface albedos 

must be specifically measured in order to accurately simulate solar flux, due to the large 

variations that may occur within a given surface type. 

Snow cover, even many kilometers from measurement sites is known to increase detected 

UV irradiances. Complicated interactions result when radiation is scattered by snow (or other 

bright surfaces) and backscattered or absorbed by atmospheric particles and clouds in the same 

vicinity (WMO, 2002). 

10.2.1.4 Data Requirements for a Surface UV-B Climatology 

A means of establishing the range of variability in UV-B at ground level for a populated 

area of interest would be to the development of a map of flux levels under typical seasonal 

conditions based on historical records. In the atmospheric sciences community, a map of this 

type is referred to as a “climatology.” 

The WMO (2002) stated that, in principle, if the spatial distribution of all UV absorbers 

and scatterers were fully known, the wavelength and angular distribution of the UV irradiance at 

the Earth’s surface could be determined with model calculations. However, the very limited 

information available on the distribution of the primary components (i.e., clouds, particles, O3, 

other UV-absorbing trace gases and surface albedo) makes detailed predictions impossible. 

Nearly all of the routine data on tropospheric O3 concentrations in the U.S. is from ground-level 

O3 monitors, such as those used to determine the attainment of the O3 air quality standards. Such 

measurements, alone, are not sufficient information for making reliable estimates of ambient O3 

concentrations above the boundary layer—information needed to establish the vertical 

distribution of O3 up to the tropopause. In an earlier assessment of the environmental effects of 

stratospheric O3 depletion (WMO, 1999), the UNEP concluded that, in view of the high spatial 

and temporal variability of surface UV radiation and the difficulty in maintaining calibration 

within networks of UV monitoring instruments, satellite-based observations are necessary to 

10-15
develop a satisfactory UV climatology. However, satellite-derived estimates of surface UV are 

limited by the availability of instruments in orbit, with available datasets comprising 

interpolations based upon a single satellite overpass per day for a given region. 

Furthermore, the current limitation in all satellite retrievals is that backscattered radiation, 

from which flux rates are derived, are limited to atmospheric layers above the surface. 

Assumptions must be made concerning aerosol and ozone absorption below the atmospheric 

limit of detection by satellites (McKenzie, et al., 2001). Above densely populated or industrial 

areas, aerosol extinctions are large and reduce UV appreciably (Gonzalez et al., 2000). 

Complete assessment of the uncertainties in predictions of UV surface flux would require 

comparisons between sparse available ground-level observations and satellite data over longer 

periods of time and for different geographical locations. No such assessment has been reported 

in the scientific literature. 

10.2.2 Factors Governing Human Exposure to Ultraviolet Radiation 

An assessment of public health benefits due to the attenuation of UV-B radiation by 

surface-level O3 requires appropriate consideration of: (1) the multiple factors that alter the flux 

of UV-B radiation at ground-level, as described above; (2) the factors that influence the extent of 

human exposure to UV-B radiation, particularly behavioral decisions; and (3) the effects of 

UV-B radiation exposure on human health. Consideration must also be given to the public 

health benefits from exposure to UV-B radiation. The present section outlines the most recent 

information on the determinants of exposure to UV-B radiation in human populations. 

Quantitative evaluation of human exposure to UV-B radiation is scientifically necessary to 

perform health risk assessment and to define subpopulations at risk for UV-B-related health 

effects. 

10.2.2.1 Outdoor Activities 

Exposure to solar UV radiation is related to one predominating factor: time spent outdoors 

during daylight hours. A large U.S. study was conducted using the EPA National Human 

Activity Pattern Survey (NHAPS) to assess UV radiation dose in Americans (Godar, 2001; 

Godar et al., 2001, 2003). The EPA NHAPS recorded the activity profiles of 9,386 Americans 

(age 0 to 60+ years) over a 24-month period to assess their exposure to various environmental 

10-16
pollutants, including UV radiation. Available UV radiation was assessed using the EPA 

UV-monitoring program. Solar radiation in the UV-A and UV-B waveband regions were 

measured daily at a monitoring site in each quadrant of the U.S. There is considerable error 

associated with quantifying UV radiation dose from exposure surveys and four UV-monitoring 

sites across the country; however, the qualitative information regarding factors that increase 

human exposure to UV radiation is still of relevance. The EPA-UV monitoring network has 

since expanded to 21 sites, located in 14 U.S. national parks and 7 urban areas across the 

United States (http://www.epa.gov/uvnet/). A UV-B monitoring network by the U.S. 

Department of Agriculture is also available for the quantitative assessment of UV radiation 

exposure (http://uvb.nrel.colostate.edu/UVB/). This monitoring network has 30+ monitoring 

sites across the United States and three additional sites in Canada and New Zealand. 

Godar et al. (2001) observed a strong seasonal preference for outdoor activities, with 

people spending the most time outdoors during the summer followed by spring, fall, and, lastly, 

winter. Because the solar erythemal (i.e., skin reddening) UV radiation dose is also highest 

during the summer, the estimated UV radiation dose of Americans was more than 10-fold greater 

in the summer compared to the winter season (Godar et al., 2001). 

Vacationing at the beach in the summer was associated with higher UV radiation exposures 

(Godar et al., 2001; Thieden et al., 2001). Even after accounting for sunscreen use at the beach, 

the erythemal UV radiation doses were more than 40% higher during a 3-week beach vacation 

compared to a 3-week stay at home (Godar et al., 2001). Danish children and adolescents were 

found to receive >50% of their annual UV radiation dose while vacationing at European beaches 

(Thieden et al., 2004a). Sunbathing also was associated with increased annual UV radiation 

dose in the Canadian National Survey on Sun Exposure and Protective Behaviours (Shoveller 

et al., 1998). Among the 3,449 adults (age 25+ years) who completed the telephone household 

survey, 21% stated that they spent time actively sunbathing. In a Danish study with 164 

participants, all children (age 1 to 12 years) and teenagers (age 13 to 19 years) as well as 94% of 

adults (age 20 to 76 years) had days with risk behavior (Thieden et al., 2004b). Teenagers, who 

had the highest number of risk-behaviors days, were found to have the highest annual UV 

radiation doses. Among teenagers, 76% (95% CI: 41, 98) of their UV radiation dose during the 

measurement period was received on risk-behavior days, as determined using personal electronic 

UV dosimeters and exposure diaries (Thieden et al., 2004b). 

10-17
An Australian study examining time profiles of daily UV radiation exposure among 

8th grade students observed that up to 47% of the daily UV radiation dose fell within the time 

periods when students were outdoors during school hours, sitting under shaded structures during 

lunch breaks and participating in routine outdoors or sports activities (Moise et al., 1999). 

Other studies also have found that participation in outdoor sports (e.g., basketball, soccer, 

golfing, swimming, cycling) significantly increased UV radiation exposure (Moehrle, 2001; 

Moehrle et al., 2000; Thieden et al., 2004a,b). 

10.2.2.2 Occupation 

Of the various factors that affect human exposure to UV radiation, occupation is also 

important. Approximately 5% of the American workforce work outdoors, as determined by the 

EPA NHAPS (Godar et al., 2001). On average, American indoor workers spend ~10% of their 

day outdoors. During their time outdoors, they are exposed to ~30% of the total ground-level 

UV flux, as measured by the EPA UV-monitoring program (Godar et al., 2001). Compared to 

indoor or in-home workers, outdoor workers are exposed to much higher levels of UV radiation 

(Kimlin et al., 1998a; Thieden et al., 2004a), frequently at levels that are above current exposure 

limits set by the International Commission on Non-Ionizing Radiation Protection (ICNIRP, 

2004). For example, Thieden et al. (2004a) observed that the annual UV radiation dose, 

estimated using personal electronic UV dosimeters and exposure diaries, was ~70% higher for 

gardeners than indoor workers. The gardeners received the majority (55%) of their UV radiation 

dose on working days (Thieden et al., 2004a). Another study found that outdoor workers 

received three to four times the annual UV radiation exposure of indoor workers (Diffey, 1990). 

At-risk working populations include farmers (Airey et al., 1997; Schenker et al., 2002), 

fishermen (Rosenthal et al., 1988), landscapers (Rosenthal et al., 1988), building and 

construction workers (Gies and Wright, 2003), physical education teachers (Vishvakarman et al., 

2001), mail delivery personnel (Vishvakarman et al., 2001), and various other workers who 

spend the majority of their day outdoors during peak UV radiation hours. 

10.2.2.3 Age 

Age may be a factor that influences human exposure to UV radiation. In a large U.S. study 

using the EPA NHAPS, the average UV radiation dose among American children (age 

10-18
<12 years) was estimated to be slightly higher (~20%) than that of adolescents (age 13 to 

19 years) (Godar, 2001). A large Canadian survey found that 89% of children (age <12 years) 

had 30 minutes or more of daily UV exposure compared to 51% for both adults (age 25+ years) 

and youth (age 15 to 24 years) (Lovato et al., 1998a, 1998b; Shoveller et al., 1998). In an 

English study (Diffey et al., 1996), UV radiation exposure was estimated in 180 children (age 9 

to 10 years) and adolescents (age 14 to 15 years) using personal film badges and exposure 

records. Once again, children were found to have received higher UV radiation exposure 

compared to adolescents (Diffey et al., 1996). However, as discussed earlier, a Danish study 

found that the annual UV radiation dose in teenagers (age 13 to 19 years) was 14-24% higher 

compared to children (age 1 to 12 years) and adults (Thieden et al., 2004b). This increase in UV 

radiation dose in the Danish teenagers was attributed to their increased risk-behavior days. 

Therefore, age may affect human exposure to UV radiation by influencing other factors of 

exposure, such as outdoor activity and risk behavior. 

Two studies examined lifetime UV radiation exposure among persons in the U.S. (Godar 

et al., 2003) and Denmark (Thieden et al., 2004b). Both studies observed that while there are 

slight differences in UV radiation dose by age, generally people receive fairly consistent UV 

doses at different age intervals throughout their lives. 

10.2.2.4 Gender 

Studies have indicated that females generally spend less time outdoors and, consequently, 

have lower UV radiation exposure compared to males (Gies et al., 1998; Godar et al., 2001; 

Shoveller et al., 1998). The U.S. study by Godar et al. (2001) observed that while both males 

and females had relatively consistent erythemal UV radiation doses throughout their lives, males 

consistently received higher overall UV doses compared to females at all age groups. Among all 

Americans, the lowest exposure to UV radiation was received in females during their childraising years (age 22 to 40 years) (Godar et al., 2001). The highest exposure was observed in 

males aged 41 to 59 years in the U.S. study (Godar et al., 2001). A similar Canadian survey 

found that younger adult males had the greatest exposures to UV radiation (Shoveller et al., 

1998). 

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10.2.2.5 Geography 

In the U.S. study by Godar et al. (2001), erythemal UV radiation doses were examined in 

persons living in northern and southern regions. Northerners and southerners were found to 

spend an equal amount of time outdoors; however, the higher solar flux at lower latitudes 

significantly increased the annual UV radiation dose for southerners (Godar et al., 2001). The 

annual UV radiation doses in southerners were 25 and 40% higher in females and males, 

respectively, compared to northerners (Godar et al., 2001). Other studies also have shown that 

altitude and latitude influence personal exposure to UV radiation (Kimlin et al., 1998b; Rigel 

et al., 1999). 

10.2.2.6 Protective Behavior 

Protective behaviors such as using sunscreen (e.g., Nole and Johnson, 2004), wearing 

protective clothing (e.g., Rosenthal et al., 1988; Sarkar, 2004; Wong et al., 1996), and spending 

time in shaded areas (Moise et al., 1999; Parisi et al., 1999) have been shown to reduce exposure 

to UV radiation. In one study, the use of sunscreen was associated with extended intentional UV 

radiation exposure (Autier et al., 1999); however, a follow-up study indicated that sunscreen use 

increased duration of exposures to doses of UV radiation that were below the threshold level for 

erythema (Autier et al., 2000). 

In a national study of U.S. youths aged 11 to 18 years, the most prevalent protective 

behavior was sunscreen use (39.2%) followed by use of a baseball hat (4.5%) (Davis et al., 

2002). There were significant differences in the use of sunscreen by age group and gender, with 

the younger age group (age 11 to 13 years) and girls having greater likelihood (47.4 and 48.4%, 

respectively) of using sunscreen (Davis et al., 2002). The Canadian National Survey on Sun 

Exposure and Protective Behaviours observed that less than half of the adults (age 25+ years, 

n = 3,449) surveyed took adequate protective actions (Shoveller et al., 1998). Once again, 

children (age <12 years, n = 1,051) were most protected from exposure to UV radiation, with 

76% using sunscreen and 36% avoiding the sun, as reported by their parents (Lovato et al., 

1998a). However, the protection level was still not adequate, as indicated by the high 45% rate 

of erythema in children. Among Canadian youth (age 15 to 24 years, n = 574), protective 

actions from UV radiation exposure included wearing a hat (38%) and seeking shade and 

avoiding the sun between the peak hours of 11:00 a.m. to 4:00 p.m. (26%) (Lovato et al., 1998b). 

10-20
The lowest prevalence of protective behavior among the youth was likely responsible for the 

highest proportion of erythema (68%) experienced in this age group. A Danish study observed 

that both children and teenagers applied sunscreen on more days than adults, but teenagers had 

the most days with erythema, due to their increased risk behavior (Thieden et al., 2004b). 

A survey in Switzerland of 1,285 individuals, including children and parents, indicated that 

sunscreen use was the protective action most commonly used, but only at the beach and not in 

routine daily exposure (Berret et al., 2002). In general, protective clothing and avoiding the sun 

were not highly used among these individuals to protect against UV-related health effects. 

10.2.2.7 Summary of Factors that Affect Human Exposures to Ultraviolet Radiation 

The factors that potentially influence UV radiation doses were discussed in the previous 

sections and include outdoor activities, occupation, age, gender, geography, and protective 

behavior. Results from the various studies indicate that the following subpopulations may be at 

risk for higher exposures to UV radiation: 

C Individuals who engage in high-risk behavior, viz., sunbathing; 

C Individuals who participate in outdoor sports and activities; 

C Individuals who work outdoors with inadequate shade, e.g., farmers, fishermen, 

gardeners, landscapers, building and construction workers; and 

C Individuals living in geographic areas with higher solar flux (i.e., lower latitudes 

[e.g., Honolulu, HI] and higher altitudes [e.g., Denver, CO]). 

10.2.3 Factors Governing Human Health Effects due to Ultraviolet Radiation 

Ultraviolet radiation occupies a specific region of the electromagnetic spectrum of 

wavelengths and can be further subdivided into three parts, UV-A (320 to 400 nm), UV-B 

(280 to 320 nm), and UV-C (200 to 280 nm). Most of the health risks associated with UV 

radiation exposure are wavelength dependent. Wavelengths <180 nm are of little practical 

biological significance as they are almost completely absorbed by the stratosphere (ICNIRP, 

2004). 

“Action spectra” of a given biological response to UV radiation across its spectral range 

are used to estimate exposure by weighting individual wavelength intensities according to the 

10-21
associated response. The overall effectiveness of the incident flux at inducing the biological

response of interest is computed by means of the relationship: 

(10-3)

where I8 and E8 are, respectively, the irradiance and its relative effectiveness at wavelength 8. 

Until 1980, it was generally thought that wavelengths <315 nm were responsible for the 

most significant adverse UV radiation health effects; however, recent studies have found that the 

longer wavelengths in the UV-A range also may produce adverse responses at substantially 

higher doses (ICNIRP, 2004). As UV-A radiation is not absorbed by O3, health effects solely 

induced by UV-A exposure are not relevant in an analysis of public health risks/benefits 

associated with O3-related UV attenuation. Therefore, this section focuses on the latest available 

information on the various adverse health effects associated with acute and chronic UV-B 

radiation exposure. 

10.2.3.1 Erythema 

Association Between Ultraviolet Radiation Exposure and Erythema 

The most conspicuous and well-recognized acute response to UV radiation is erythema, or 

the reddening of the skin, which is likely caused by direct damage to DNA by UV-B and UV-A 

radiation (Matsumura and Ananthaswamy, 2004). Indirect oxidative damage also may occur at 

longer wavelengths (Matsumura and Ananthaswamy, 2004). Skin type appears to play a large 

role in the sensitivity to UV radiation-induced erythema. The Fitzpatrick classifications for skin 

types are: (1) skin type I – individuals with extremely sensitive skin that sunburns easily and 

severely, and is not likely to tan (e.g., very fair skin, blue eyes, freckles); (2) skin type 

II – individuals with very sensitive skin that usually sunburns easily and severely, and tans 

minimally (e.g., fair skin, red or blond hair, blue, hazel or brown eyes); (3) skin type 

III – individuals with sensitive skin that sunburns moderately and tans slowly (e.g., white skin, 

dark hair); (4) skin type IV – individuals with moderately sensitive skin that sunburns minimally 

and usually tans well (e.g., white or light brown skin, dark hair, dark eyes); (5) skin type 

V – individuals with minimally sensitive skin that rarely sunburns and tans deeply (e.g., brown 

skin); and (6) skin type VI – individuals with nonsensitive skin that never sunburns and tans 

10-22
profusely (e.g., dark skin). Harrison and Young (2002) found that the perceptible minimal 

erythemal dose was approximately twofold greater for individuals with skin type IV compared to 

skin type I, although there was considerable overlap in the minimal erythemal dose among the 

four skin types. Waterston et al. (2004) further observed that within an individual, erythemal 

response differed by body site (e.g., abdomen, chest, front upper arm, back of thigh). These 

differences were likely attributable to body site-specific variations in melanin pigmentation. 

Kollias et al. (2001) investigated the change in erythemal response following a previous 

exposure to UV radiation. Body sites that received a second exposure to UV radiation always 

showed a reduced erythemal response compared to body sites with a single exposure, especially 

when the first exposure was at levels greater than the minimal erythemal dose. The suppression 

of erythema was more pronounced when the second exposure was given 48 hours after the first. 

These findings support the well established notion that repeated exposures to UV radiation 

results in adaptation (e.g., stimulation of melanogenesis). Kaidbey and Kligman (1981) 

examined individuals with skin types I, II, and III, and found that multiple exposures to 

subthreshold doses of UV radiation at 24-hour intervals resulted in cumulative injury to the skin, 

as indicated by a lowering of the minimal erythemal dose. These results suggest that a longer 

time period than 24 hours may be necessary to repair damage from a single exposure to UV 

radiation. Henriksen et al. (2004) also observed a lowering of the minimal erythemal dose with 

repeated exposure at 24-hour intervals in 49 healthy volunteers with skin types II, III, IV, and V. 

However, adaptation was reached after the 4th consecutive exposure. Henriksen et al. further 

found that the change in threshold depended on skin type. After 4 days of repeated UV 

radiation, there was little change (10 to 20%) in the erythemal threshold dose with repeated 

exposure to UV radiation in the fair-skinned individuals. Among the darker-skinned individuals, 

the minimal erythemal dose was lowered by 40 to 50%. However, both the initial UV dose and 

the dose to erythema after four days of exposure was still higher in the dark-skinned persons. 

A reference erythema action spectrum was adopted by the Commission Internationale de 

l’Eclairage (International Commission on Illumination, CIE) in 1987 (McKinlay and Diffey, 

1987). The CIE erythema action spectrum indicates that UV-B radiation is orders of magnitude 

more effective per unit dose than UV-A radiation. 

10-23
Risk of Erythema from Changes in Tropospheric O3 Levels 

There is no literature examining the risk of erythema associated with changes specifically 

in tropospheric or ground-level O3 levels. The scientific studies, available to date, focus on the 

effects of a reduction in stratospheric ozone. One such study has assessed the effects of 

stratospheric O3 depletion on the risk of erythema (Longstreth et al., 1998). The analysis by 

Longstreth et al. (1998) concluded that the risk of erythema would not appreciably increase with 

depletion of the stratospheric O3 layer. This is due to the powerful adaptation of the skin to 

different levels of UV radiation, as evidenced by its ability to cope with changes in UV radiation 

by season (Van Der Leun and De Gruijl, 1993). Gradual exposure to increasing UV radiation 

from the winter to summer leads to decreased sensitivity of the skin. In midlatitudes, the UV-B 

radiation in the summer is 10-fold greater than in the winter. In contrast, the steady depletion of 

the O3 layer has been estimated to result in an approximately 20% increase in UV-B over 10 

years (Longstreth et al., 1998). The comparatively small increase in UV radiation throughout the 

years, therefore, would not significantly increase the risk of erythema. Tropospheric O3 

constitutes no more than 10% of total atmospheric O3. Given that stratospheric O3 depletion was 

unlikely to increase the risk of erythema, one could reasonably conclude that small changes in 

ground-level O3 that take place with attainment of the O3 NAAQS would also not result in 

increased risk. 

10.2.3.2 Skin Cancer 

According to the American Academy of Dermatology, one in five Americans develop skin 

cancer during their lifetime. The three main forms of skin cancer include basal cell carcinoma 

and squamous cell carcinoma, which are both nonmelanoma skin cancers, and malignant 

melanoma. Nonmelanoma skin cancers constitute more than one-third of all cancers in the U.S. 

and ~90% of all skin cancers, with basal cell carcinoma being approximately four times as 

common as squamous cell carcinoma (Diepgen and Mahler, 2002; ICNIRP, 2004). The 

incidence of malignant melanoma is much lower than nonmelanoma skin cancers. In 2004, 

more than one million cases of basal and squamous cell skin cancer are expected to be newly 

diagnosed, compared to 40,780 cases of melanoma (Jemal et al., 2004). However, melanoma 

has great metastatic potential and accounts for the majority of skin cancer deaths. 

10-24
Exposure to UV radiation is considered to be a major risk factor for all three forms of skin 

cancer (Gloster and Brodland,1996; Diepgen and Mahler, 2002; IARC, 1992). Ultraviolet 

radiation is especially effective in inducing genetic mutations and acts as both a tumor initiator 

and promoter. Keratinocytes have evolved DNA repair mechanisms to correct the damage 

induced by UV; however, mutations can occur, leading to skin cancers that are appearing with 

increasing frequency (Hildesheim and Fornace, 2004). The relationship between skin cancer and 

chronic exposure to UV radiation is further explored below, followed by discussion of the 

influence of O3 on the incidence of skin cancer. 

10.2.3.3 Ultraviolet Radiation Exposure and the Incidence of Nonmelanoma Skin Cancers 

The incidence of all three types of cancers has been shown to rise with increasing UV 

radiation concentrations across the U.S. (De Gruijl, 1999); however, the most convincing 

evidence for a causal relationship exists between UV radiation and squamous cell carcinoma. 

Squamous cell carcinoma occurs almost exclusively on skin that is regularly exposed to the sun, 

such as on the face, neck, arms, and hands. The incidence is higher among whites in areas of 

lower latitudes, where solar flux is greater (Kricker et al., 1994). The risk of squamous cell 

carcinoma was shown to increase with life-long accumulated exposure to UV radiation in one 

cross-sectional study (Vitasa et al., 1990); however, increased risk was found to be associated 

only with exposure 10 years prior to diagnosis in a case-control study (Gallagher et al., 1995a). 

One of the major concerns with both types of studies is the potential for recall bias in reporting 

past UV radiation exposure by individuals already aware of their disease status. 

Ultraviolet radiation also has been linked to basal cell carcinoma. Basal cell carcinoma is 

common on the face and neck (80-90%) but rarely occurs on the back of the hands (De Gruijl, 

1999). While cumulative UV radiation exposure was not associated with an increased risk of 

basal cell carcinoma (Vitasa et al., 1990), increased risk was observed in individuals with greater 

recreational UV radiation exposure in adolescence and childhood (age <19 years) and 

individuals with a history of severe erythema in childhood (Gallagher et al., 1995b). Once again, 

consideration must be given to potential recall bias in assessing these results. Thus, there is 

suggestive evidence that UV radiation also plays a role in the development of basal cell 

carcinoma, but the etiologic mechanisms for squamous cell carcinoma and basal cell carcinoma 

likely differ. In an Australian study conducted in a subtropical community, the factors of having 

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fair skin, a history of repeated sunburns, and nonmalignant solar skin damage diagnosed by 

dermatologists were strongly associated with both types of nonmelanoma skin cancer (Green 

et al., 1996). The authors attributed the finding that outdoor occupation was not associated with 

nonmelanoma skin cancer to self-selection. Individuals with fair or medium complexions and a 

tendency to sunburn accounted for more than 80% of the community study sample; however, 

they were systematically underrepresented among outdoor workers (Green et al., 1996). Such 

self-selection bias might partly explain the lack of consistent quantitative evidence of a causal 

link between UV radiation and skin cancer in humans. 

De Gruijl et al. (1993) assessed the action spectrum for nonmelanoma skin cancers using 

hairless albino mice. Human data are not available regarding wavelength dependence of the 

carcinogenicity of UV radiation. After adjusting for species differences, the Skin Cancer 

Utrecht-Philadelphia action spectrum indicated the highest effectiveness in the UV-B range with 

a maximum at 293 nm, which dropped to 10!4

 of this maximum at the UV-A range above 

340 nm (De Gruijl et al., 1993). The mutations commonly present in the p53 tumor suppressor 

gene in individuals with squamous cell carcinoma and basal cell carcinoma are called the 

“signature” mutations of UV-B radiation (De Gruijl, 2002). UV-B radiation is highly mutagenic, 

because DNA is a chromophore for UV-B, but not for UV-A radiation (Ichihashi et al., 2003). 

Nevertheless, other studies have found that UV-A radiation, in addition to UV-B radiation, can 

induce DNA damage (Persson et al, 2002; Rünger et al., 2000). DNA damage by UV-A is 

mediated by reactive oxygen species, making it indistinguishable from damage caused by other 

agents that generate reactive oxygen species (De Gruijl, 2002). Epidemiologic evidence of a 

carcinogenic effect of UV-A was found in a study of psoriasis patients receiving oral psoralen 

and UV-A radiation treatment (Stern et al., 1998). High-dose exposure to oral psoralen and 

UV-A radiation was associated with a persistent, dose-related increase in the risk of squamous 

cell cancer. Risk of basal cell cancer also was increased in those patients exposed to very high 

levels of UV-A radiation. Therefore, although UV-B radiation has long been considered the 

main culprit for nonmelanoma skin cancer, limited evidence suggest that UV-A radiation may 

also play a role. 

Susceptible populations for nonmelanoma skin cancers include individuals with reduced 

capacity for nucleotide excision repair, the primary repair mechanism for UV radiation-induced 

DNA lesions (Ichihashi et al., 2003). At particular risk are individuals with xeroderma 

10-26
pigmentosum, as they have defective nucleotide excision repair in all tissues (Kraemer, 1997; 

Sarasin, 1999). Skin type also largely affects susceptibility to skin cancer. Of the six skin 

phenotypes, the most sensitive individuals are those with skin types I and II, who have a fair 

complexion, blue or green eyes, and red or blond hair (Diepgen and Mahler, 2002). These 

individuals tend to sunburn easily, tan poorly, and freckle with sun exposure. A history of 

repeated sunburns also appears to increase the risk of both cancers, while sunburns during 

childhood are more associated with increased basal cell carcinoma (Gallagher et al., 1995b; 

Green et al., 1996). 

Ultraviolet Radiation and the Incidence of Cutaneous Malignant Melanoma 

From 1973 to 1994, the incidence rate of melanoma increased 120.5% along with an 

increased mortality rate of 38.9% among whites in the United States (Hall et al., 1999). The 

ICNIRP (2004) states that during the past 40 years or so, each decade has seen a twofold 

increase in the incidence of malignant melanoma in white populations, with increased incidence 

observed more prominently in individuals living in lower latitudes. Cutaneous malignant 

melanoma has a mutifactorial etiology with environmental, genetic, and host factors (Lens and 

Dawes, 2004). The major environmental factor of malignant melanoma has been identified as 

UV radiation exposure (Diepgen and Mahler, 2002); therefore, the increased incidence of 

melanoma throughout the years might be partially attributable to changes in human activity 

patterns (e.g., increased outdoor activity) that influence UV exposure or increased UV radiation 

at the ground level. The risk of melanoma appears to depend on the interaction between the 

nature of the exposure and skin type (Lens and Dawes, 2004). 

Fears et al. (2002) examined the association between invasive cutaneous melanoma and 

UV radiation in non-Hispanic whites using a case-control study design. Lifetime residential 

history was coupled with mid-range UV-B radiation flux measurements to reduce exposure 

misclassification and recall bias. A 10% increase in the average annual UV-B flux was 

significantly associated with a 19% (95% CI: 5, 35) increase in individual odds for melanoma in 

men and a 16% (95% CI: 2, 32) increase in women. Whiteman et al. (2001) conducted a 

systematic review of studies that examined the association between childhood UV radiation 

exposure and risk of melanoma. Researchers found that ecological studies assessing ambient 

sun exposure consistently reported higher risks of melanoma among people who resided in an 

10-27
environment with high UV radiation during their childhood (Whiteman et al., 2001). The lack of 

consistency among the case-control studies was likely due to the varying methods used to assess 

UV radiation dose. 

While the evidence is generally suggestive of a causal relationship between UV radiation 

and malignant melanoma, possibly conflicting data also has been observed. For example, the 

highest occurrence of malignant melanoma is on men’s backs and women’s legs, areas that do 

not have prolonged exposure to the sun (Rivers, 2004). This indicates that, unlike nonmelamona 

skin cancers, malignant melanoma tends to occur in sites of intermittent, intense sun exposure 

(trunk and legs), rather than in areas of cumulative sun damage (head, neck, and arms) (Swetter, 

2003). A study by Whiteman et al. (2003) observed that individuals with melanomas of the 

trunk had more melanocyte nevi and less solar keratoses compared to individuals with head and 

neck melanomas, suggesting that cutaneous melanomas may arise through two pathways, one 

associated with melanocyte proliferation and the other with chronic exposure to sunlight. Green 

et al. (1999) also found that melanomas of the soles and palms resembled other cutaneous 

melanomas in their association with sun exposure, but were distinguished from them by their 

strong positive associations with nevi on the soles. 

The available data conflict with regard to the relative importance of UV-A versus UV-B 

in inducing melanomas. UV-A has a much higher flux rate at the Earth’s surface, as it is not 

absorbed by O3; and it is able to penetrate more deeply into the skin surface due to its longer 

wavelength. However, UV-B, as mentioned earlier, is much more energetic and, therefore, more 

effective in photochemically altering DNA. The individual roles of UV-A and UV-B in the 

development of cutaneous malignant melanoma have been examined in several studies. 

A case-control study of 571 patients and 913 matched controls found an elevated odds ratio of 

1.8 (95% CI: 1.2, 2.7), after adjusting for skin type, hair color, raised nevi, and number of 

sunburns, for developing malignant melanoma in individuals who regularly used tanning beds, 

which typically are UV-A sources (Westerdahl et al., 2000). In a study by Setlow et al. (1993), 

an action spectrum using the tropical fish Xiphophorus indicated that UV-A range wavelengths 

were especially important in malignant melanoma induction. However, an action spectrum 

using the opossum Monodelphis domestica found that the potency of UV-A for melanoma 

induction was extremely low compared to that of UV-B (Robinson et al., 2000). A recent study 

by De Fabo et al. (2004) examined the differences in wavelength effectiveness using a 

10-28
hepatocyte growth factor/scatter factor-transgenic mouse model. The epidermal tissue of these 

transgenic mice behaves similar to the human epidermis in response to UV exposure. Given the 

absence of a mammalian melanoma action spectrum, the standardized CIE erythema action 

spectrum was used to deliver identical erythemally effective doses. Only UV-B radiation was 

found to initiate mammalian cutaneous malignant melanoma. UV-A radiation, even at doses 

considered physiologically relevant, were ineffective at inducing melanoma (De Fabo et al., 

2004). Overall, current evidence suggests that UV-B, and not UV-A, is the primary risk factor 

for malignant melanoma (ICNIRP, 2004). 

The populations susceptible for malignant melanoma are similar to those for nonmelanoma 

skin cancers. Once again, individuals with xeroderma pigmentosum or a reduced capacity for 

nucleotide excision repair are at increased risk (Tomescu et al., 2001; Wei et al., 2003). 

Individuals with skin types I and II, or the fair-skin phenotype (blue or green eyes; blond or red 

hair; skin that freckles, sunburns easily, and does not tan), have increased susceptibility to 

malignant melanoma (Evans et al., 1988; Swetter, 2003; Veierød et al., 2003). However, the 

incidence of melanoma was also positively associated with UV radiation in Hispanics and blacks 

(Hu et al., 2004). Although the incidence of melanoma is much lower in Hispanics and blacks 

compared to whites, melanomas in these populations are more likely to metastasize and have a 

poorer prognosis (Black et al., 1987; Bellows et al., 2001). Among children, malignant 

melanoma appears to have similar epidemiologic characteristics to the adult form of the disease 

(Whiteman et al., 1997). Individuals with intermittent, intense sun exposure, particularly during 

childhood, were found to have increased risk of melanoma (Whiteman et al., 2001), in contrast 

to the association between cumulative exposure and increased risk of squamous cell carcinoma. 

One study found that a personal history of nonmelanoma skin cancer or precancer, higher 

socioeconomic status, and increased numbers of nevocytic nevi also were associated with 

increased incidence of melanoma (Evans et al., 1988). 

Effect of Changes in Tropospheric O3 Levels on Skin Cancer Incidence 

The current evidence strongly suggests a causal link between exposure to UV radiation and 

the incidence of both nonmelanoma and melanoma skin cancer. Genetic factors, including skin 

phenotype and ability to repair DNA, affect an individual’s susceptibility to skin cancer. 

Quantifying the relationship between UV radiation and skin cancer is complicated by the 

10-29
uncertainties involved in the selection of an action spectrum and appropriate characterization of 

dose (e.g., peak or cumulative levels of exposure, childhood or lifetime exposures). In addition, 

there are multiple complexities in attempting to quantify the effect of tropospheric O3 levels on 

UV-radiation exposure, as described in Section 10.2. The absence of published studies that 

critically examine increased incidence of skin cancer attributable to decreased tropospheric O3 

levels reflects the significant challenges in determining ground-level O3-related changes in UV 

radiation exposure. An analysis by Lutter and Wolz (1997) attempted to estimate the effects of 

a nationwide 10 ppb reduction in seasonal average tropospheric O3 on the incidence of 

nonmelanoma and melanoma skin cancers and cataracts. Their estimate, however, depended 

upon several simplifying assumptions, ranging from an assumed generalized 10 ppb reduction 

in O3 column density, national annual average incidence rates for the two types of skin cancer, 

and simple, linear biological amplification factors. Further, the methodologies used in this 

analysis inherently have ignored area-specific factors that are important in estimating the extent 

to which small, variable changes in ground-level O3 mediate long-term exposures to UV-B 

radiation. More reasonable estimates of the human health impacts of enhanced UV-B 

penetration following reduced surface O3 concentrations require both (a) solid understanding 

of the multiple factors that define the extent of human exposure to UV-B at present and 

(b) well-defined and quantifiable links between human disease and UV-B exposure. The reader 

is referred to the U.S. EPA 2002 Final Response to Court Remand (Federal Register, 2003) for 

detailed discussions of the data and scientific issues associated with the determination of public 

health benefits resulting from the attenuation of UV-B by surface-level O3. 

In the absence of studies specifically addressing the reduction of tropospheric O3 (by 

assuming that the key variable is total column O3 density), inferences could be made concerning 

the effects of reduced tropospheric O3-related increases in UV-B exposure on the basis of studies 

focused on stratospheric O3 depletion. Several studies have examined the potential effect of 

stratospheric O3 depletion on the incidence of skin cancer (De Gruijl, 1995; Longstreth et al., 

1995; Madronich and De Gruijl, 1993; Slaper et al., 1996; Urbach, 1997). Note that several of 

the concerns expressed in relation to the Lutter and Wolz (1997) analysis are relevant here as 

well. Stratospheric O3 depletion is likely to increase the ground-level UV-B flux, as O3 absorbs 

radiation in that wavelength range with high efficiency. Because UV-B radiation is primarily 

implicated in the induction of skin cancer, especially among persons with skin phenotypes I and 

10-30
II, there is concern that the depletion of the O3 layer would result in significantly increased 

incidence of skin cancers. 

Estimation of the increased risk in melanoma associated with stratospheric O3 depletion 

cannot be done adequately due to the lack of a mammalian action spectrum for melanoma. 

Furthermore, the complexity of the UV-related induction mechanism of melanoma adds an 

additional layer of uncertainty to the calculations. The excess risk in nonmelanoma skin cancers 

associated with a decrease in stratospheric O3 was estimated using the Skin Cancer UtrechtPhiladelphia action spectrum based on hairless albino mice (Longstreth et al., 1995). 

Quantification of how much more UV radiation would reach ground level with each percentage 

decrease in O3 required several assumptions: (1) annual doses were an appropriate measure; (2) 

personal doses were proportional to ambient doses; and, most notably, (3) each percentage 

decrease in O3 was associated with a 1.2% increase in UV radiation. Next, the relationship 

between UV radiation and nonmelanoma skin cancer incidence was determined: each percent 

increase in annual UV radiation dose was estimated to cause a 2.5% increase in squamous cell 

carcinoma and 1.4% increase in basal cell carcinoma over a human lifetime. Incorporating all 

these factors, Longstreth et al. (1995) calculated that a sustained 10% decrease in stratospheric 

O3 concentration would result in 250,000 additional nonmelanoma skin cancer cases per year. 

Madronich and De Gruijl (1993) noted that the largest percent of O3-induced nonmelanoma skin 

cancer increases would be at high latitudes, where baseline incidence of skin cancer is usually 

small. Assuming a phaseout of primary O3-depleting substances by 1996, as established by the 

Copenhagen Amendments in 1992, Slaper et al. (1996) estimated that the number of excess 

nonmelanoma skin cancers in the U.S. caused by O3 depletion would exceed 33,000 per year (or 

approximately 7 per 100,000) around the year 2050. 

However, estimating the increase in nonmelanoma skin cancer incidence attributable to the 

depletion of the stratospheric O3 layer is marred by uncertainty. The following statement by 

Madronich and De Gruijl (1994) describes the uncertainty of estimating the effect of 

stratospheric O3 depletion on the incidence of skin cancer: 

Extrapolating trends and effects of UV into the future is very hypothetical due to uncertainties 

that arise from atmospheric chemistry, epidemiology, and related disciplines. The values that we 

calculated are one plausible measure of the magnitude of the O3-UV effects... The timescales for 

atmospheric change and skin-cancer development are still far from certain: O3 reductions are 

expected to continue well into next century, and the time between UV exposure and 

development of skin cancer is essentially unknown. 

10-31
Therefore, much caution is necessary when assessing and interpreting the quantitative results of 

excess nonmelanoma skin cancer incidence due to stratospheric O3 depletion. Although the 

effect of reductions in tropospheric or ground-level O3 concentrations on skin cancer incidence 

has not been assessed, it would be expected to be much less compared to the effect from the 

depletion of the stratospheric O3 layer, given that tropospheric O3 makes up #10% of the total 

atmospheric O3. 

10.2.3.4 Ocular Effects of Ultraviolet Radiation Exposure 

Ultraviolet Radiation Exposure and Risk of Ocular Damage 

Ocular damage from UV radiation exposures includes effects on the cornea, lens, iris, and 

associated epithelial and conjunctival tissues. Absorption of UV radiation differs by 

wavelength, with short wavelengths (<300 nm) being almost completely absorbed by the cornea, 

whereas longer wavelengths are transmitted through the cornea and absorbed by the lens 

(McCarty and Taylor, 2002). The most common acute ocular effect of environmental UV 

exposure is photokeratitis, also known as snowblindness, caused by absorption of short 

wavelength UV radiation by the cornea. The action spectrum indicated that maximum 

sensitivity of the human eye was found to occur at 270 nm (ICNIRP, 2004; Pitts, 1993). The 

threshold for photokeratitis in humans varied from 4 to 14 mJ/cm2

 for wavelengths 220 to 

310 nm. 

Exposure to longer wavelengths has been shown to cause both transient and permanent 

opacities of the lens, or cataracts. Extensive toxicologic and epidemiologic evidence supports 

the causal association between UV radiation and cataracts (Hockwin et al., 1999; McCarty and 

Taylor, 2002). Ultraviolet radiation-induced cataracts are hypothesized to be caused by 

oxidative stress leading to increased reactive species in the lens, which then causes damage to 

lens DNA and cross-linking of proteins. Exposure time to low-dose UV radiation was found to 

strongly influence cataract formation (Ayala et al., 2000). An action spectrum determined using 

young female rats indicated that the rat lens was most sensitive to 300 nm, correcting for corneal 

transmittance (Merriam et al., 2000). Oriowo et al. (2001) examined the action spectrum for 

cataract formation using whole cultured pig lenses. As pig lenses are similar in shape and size 

to the human lens, some inferences may be made. Results indicated that the 270 to 315 nm 

waveband was most effective in producing UV-induced cataracts in vitro. However, the 

10-32
threshold values varied widely within that range, from 0.02 J/cm2

 for 285 nm to 0.74 J/cm2 for 

315 nm (Oriowo et al., 2001). At wavelengths >325 nm, the threshold levels were orders of 

magnitude larger, with a minimum threshold value of 18.7 J/cm2

. 

An epidemiologic study examined the effects of UV radiation on cataract formation in 

watermen (e.g., commericial fishermen, boat workers) who worked on Chesapeake Bay, MD 

(Taylor et al., 1988). Among the 838 individuals surveyed in this study, 111 had cortical 

cataracts and 229 had nuclear cataracts. Results indicated that UV-B radiation was significantly 

associated with cortical, but not nuclear, cataract formation. For a given age, a doubling of 

cumulative UV-B exposure was associated with a 60% excess risk (95% CI: 1, 164) of cortical 

cataracts. No association was observed between cataracts and UV-A radiation in this outdoorworking population. 

Risk of Ocular Damage from Changes in Tropospheric Ozone Levels 

Cataracts are the most common cause of blindness in the world. McCarty et al. (2000) 

calculated that ocular UV radiation exposure accounted for 10% of the cortical cataracts in an 

Australian cohort of 4,744 individuals from both urban and rural areas. A study by Javitt and 

Taylor (1994-1995) found that the probability of cataract surgery in the U.S. increased by 3% for 

each 1° decrease in latitude. These results suggest that depletion of the stratospheric O3 layer 

may increase UV radiation-induced cataract formation. After assuming a certain wavelength 

dependency along with several additional assumptions, every 1% decrease in the stratospheric O3 

layer was estimated to be associated with a 0.3 to 0.6% increase in cataracts (Longstreth et al., 

1995). Longstreth et al. (1995) noted that this estimate has a high degree of uncertainty due to 

inadequate information on the action spectrum and dose-response relationships. Quantitative 

estimates have not been possible for photokeratitis, pterygium, or other UV-related ocular effects 

due to lack of epidemiologic and experimental data. 

As is the case for all of the other UV-related health outcomes, there is no published 

information on the potential effects on cataract formation due to any changes in surface-level 

UV flux resulting from decreases in tropospheric O3. 

10-33
10.2.3.5 Ultraviolet Radiation and Immune System Suppression 

Experimental studies have suggested that exposure to UV radiation may suppress local and 

systemic immune responses to a variety of antigens (Clydesdale et al., 2001; Garssen and Van 

Loveren, 2001; Selgrade et al., 1997). In rodent models, these effects have been shown to 

worsen the course and outcome of some infectious diseases and cancers (Granstein and Matsui, 

2004; Norval et al., 1999). Granstein and Matsui (2004) stated that exposure to UV-B radiation 

caused immunosuppression in mice ultimately by releasing cytokines that prevent antigenpresenting cells from performing their normal functions and causing direct damage to epidermal 

Langerhans cells. Noonan et al. (2003) investigated UV skin cancer induction in two strains of 

reciprocal F1 hybrid mice and found that genetically determined differences in susceptibility to 

UV-induced immunosuppression was a risk factor for skin cancer. At high UV radiation doses, 

mice with greater susceptibility to immune suppression had a larger proportion of skin tumors 

compared to those with lower susceptibility (Noonan et al., 2003). In a study by Yoshikawa 

et al. (1990), development of contact hypersensitivity to dinitrochlorobenzene on irradiated 

buttock skin was examined. Individuals who failed to develop contact hypersensitivity were 

considered to be susceptible to UV-B radiation. Virtually all skin cancer patients (92%) were 

susceptible to UV-B radiation-induced suppression of contact hypersensitivity, compared to 

approximately 40% of healthy volunteers. Others studies have observed increased skin cancer in 

immune suppressed organ transplant patients (Caforio et al., 2000; Lindelöf et al., 2000). 

Collectively, results from these studies suggest that immune suppression induced by UV 

radiation may be a risk factor contributing to skin cancer induction (Ullrich, 2005). 

There is also some evidence that UV radiation has indirect involvement in viral 

oncogenesis through the human papillomavirus (Pfister, 2003). Additional evidence of 

UV-related immunosuppression comes from an epidemiologic study of 919 patients with rare 

autoimmune muscle diseases from 15 cities on four continents with variable UV radiation 

intensity (Okada et al., 2003). Ultraviolet radiation was strongly associated with the prevalence 

of dermatomyositis, an autoimmune disease distinguished by the presence of photosensitive 

pathognomonic rashes (Okada et al., 2003). In patients with the human immunodeficiency virus, 

UV-B radiation lead to activation of the virus in their skin through the release of cytoplasmic 

nuclear factor kappa B (Breuher-McHam et al., 2001). In a study by Selgrade et al. (2001), 

UV-induced immunosuppression was examined in 185 subjects with different skin 

10-34
pigmentations. To assess immune suppression, dinitrochlorobenzene was applied to irradiated 

buttock skin 72 hours after irradiation. Differences in sensitivity were unrelated to skin type 

based on the Fitzpatrick classification or minimal erythemal dose (Selgrade et al., 2001). 

However, erythemal reactivity, assessed by the steepness of the erythemal dose-response curve, 

was shown to be significantly associated with UV-induced immunosuppression. Only subjects 

with steep erythemal responses, which included individuals with skin types I through V, showed 

a dose-response relationship between UV exposure and immune suppression (Selgrade et al., 

2001). 

In other studies, UV radiation was associated with decreased autoimmune diseases. 

Several ecologic studies observed a decreased prevalence of multiple sclerosis, insulindependent diabetes mellitus, and rheumatoid arthritis in regions with lower latitude (i.e., higher 

UV radiation exposure) (Ponsonby et al., 2002). These results may be attributable to UV 

radiation-induced immunosuppression and UV-B-related production of vitamin D, which has 

immunomodulatory effects (Cantorna, 2000). The protective effects of UV radiation resulting 

from its active role in vitamin D production are further discussed in the next section. 

Most action spectrum investigations have concluded that immunosuppression is caused 

most effectively by the UV-B waveband (Garssen and Van Loveren, 2001). The effects of UVA on local and systemic immunosuppression have been unclear and inconsistent. There is some 

evidence that high doses of UV-A are protective of immunosuppression induced by UV-B 

exposure (Halliday et al., 2004). Given the variety of outcomes of immune suppression and 

possible mechanisms of effect, little detailed information exists on UV radiation action spectra 

and dose-response relationships. The available data are insufficient to conduct a critical risk 

assessment of UV radiation-induced immunosuppression in humans. 

10.2.3.6 Protective Effects of Ultraviolet Radiation—Production of Vitamin D 

Any risk assessment that attempts to quantify the consequences of increased UV-B 

exposure on humans due to reduced ground-level O3 must include consideration of both negative 

and positive effects. A potential health benefit of increased UV-B exposure relates to the 

production of vitamin D in humans. Most humans depend on sun exposure to satisfy their 

requirements for vitamin D (Holick, 2004). UV-B photons are absorbed by 

7-dehydrocholesterol in the skin, leading to its transformation to previtamin D3, which is rapidly 

10-35
converted to vitamin D3. Vitamin D3 is metabolized in the liver, then in the kidney to its 

biologically active form of 1.25-dihydroxyvitamin D3. One minimal erythemal dose produces 

vitamin D equivalent to an oral dose of 20,000 IU vitamin D, which is 100 times the 

recommended dietary allowance for adults under 50 years of age (Giovannucci, 2005; 

Holick, 2004). 

Vitamin D deficiency can cause metabolic bone disease among children and adults, and 

also may increase the risk of many common chronic diseases, including type I diabetes mellitus 

and rheumatoid arthritis (Holick, 2004). Substantial in vitro and toxicologic evidence also 

support a role for vitamin D activity against the incidence or progression of various cancers 

(Giovannucci, 2005; Studzinski and Moore, 1995). Large geographical gradients in mortality 

rates for a number of cancers in the U.S. are not explained by dietary or other risk factors; 

therefore, it has been hypothesized that some carcinomas may be due to latitude-related 

reduction in UV-B radiation exposure. Published literature indicates that solar UV-B radiation, 

by increasing vitamin D production, is associated with a reduced risk of cancer. Most of these 

studies used an ecologic study design, in which latitude gradient was examined in relation to 

cancer rates. Kimlin and Schallhorn (2004) observed that latitude was a valid predictor of 

vitamin D-producing UV radiation. The strongest evidence exists for an association between 

UV radiation and reduced risk of colorectal cancer (Giovannucci, 2005; Grant and Garland, 

2004; Freedman et al., 2002). Several other studies also have found an inverse relationship 

between UV radiation and various other cancers, including cancer of the breast (Freedman et al., 

2002; Garland et al., 1990; Gorham et al., 1990; Grant, 2002a; John et al., 1999), ovary 

(Freedman et al., 2002; Lefkowitz and Garland, 1994), and prostate (Freedman et al., 2002; 

Hanchette and Schwartz, 1992), as well as non-Hodgkin lymphoma (Hughes et al, 2004; Hartge 

et al., 1996). Eight other cancers (i.e., bladder, esophageal, kidney, lung, pancreatic, rectal, 

stomach, and corpus uteri) have been found to exhibit an inverse correlation between mortality 

rates and UV-B radiation (Grant, 2002b). 

Using UV-B data from July 1992 and U.S. cancer mortality rates from 1970 to 1994, 

premature cancer deaths attributable to insufficient UV-B exposure were analyzed in an ecologic 

study (Grant, 2002b). The minimum mortality rate, which was determined as the value 

corresponding to the maximum UV-B dose, was used to calculate the number of premature 

deaths. This analysis observed that the annual number of premature deaths from various cancers 

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due to latitude-related reduction in UV-B exposure was 21,700 (95% CI: 20,400, 23,400) for 

white Americans; 1,400 (95% CI: 1,100, 1,600) for black Americans; and 500 (95% CI: 400, 

600) for Asian Americans and other minorities. Uncertainty in the estimations of UV-B 

exposure limits the confidence for the estimates of excess cancer deaths attributable to 

insufficient exposure. Caution is required in interpreting results from ecologic data; however, no 

strong alternative explanation is indicated in the association observed between UV radiation and 

the decreased risk of cancer (Giovannucci, 2005). No study has assessed the decreased risk of 

cancer mortality resulting from increased UV radiation attributable to decreased tropospheric O3 

levels, but the change in risk is expected to be unappreciable. 

In establishing guidelines on limits of exposure to UV radiation, the ICNIRP agreed that 

some low-level exposure to UV radiation has health benefits (ICNIRP, 2004). However, the 

adverse health effects of higher UV exposures necessitated the development of exposure limits 

for UV radiation. The ICNIRP recognized the challenge in establishing exposure limits that 

would achieve a realistic balance between beneficial and adverse health effects. 

10.2.4 Summary and Conclusions for Ozone Effects on UV-B Flux 

Latitude and altitude are primary variables in defining UV-B flux at the Earth’s surface, 

immediately followed in importance by clouds, surface albedo, PM concentration and 

composition, and then by gas phase pollution. Of all of these, only latitude and altitude can be 

defined with small uncertainty in any effort to develop a UV climatology for use in a public 

health benefits analysis relevant to the areas not presently attaining the NAAQS for O3. Cloud 

cover, and its effect on surface UV flux, continues to be extremely difficult to define and predict. 

Particulate matter and gas-phase tropospheric pollutants are subject to similarly high degrees of 

uncertainty in predicting their relative concentration distributions. Land cover and, 

consequently, surface albedos are highly variable at the geographic scales relevant to NAAQS 

attainment. 

Within the uncertain context of presently available information on UV-B surface fluxes, a 

risk assessment of UV-B-related health effects would need to factor in human habits (e.g., daily 

activities, recreation, dress, and skin care) in order to adequately estimate UV-B exposure levels. 

Little is known about the impact of variability in these human factors on individual exposure to 

UV radiation. Furthermore, detailed information does not exist regarding the relevant type (e.g., 

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peak or cumulative) and time period (e.g., childhood, lifetime, or current) of exposure, 

wavelength dependency of biological responses, and interindividual variability in UV resistance. 

Recent reports of the necessity of UV-B in the production of vitamin D—a vitamin in which 

many individuals are deficient—suggests that increased risks of human disease due to a slight 

excess in UV-B radiation exposure may be offset by the benefits of enhanced vitamin D 

production. However, as with other impacts of UV-B on human health, this beneficial effect of 

UV-B has not been studied in sufficient detail to allow for a credible health benefits assessment. 

In conclusion, the effect of changes in surface-level O3 concentrations on UV-induced health 

outcomes cannot yet be critically assessed within reasonable uncertainty. 

10.3 TROPOSPHERIC OZONE AND CLIMATE CHANGE 

Water vapor, CO2, O3, N2O, CH4, CFCs, and other polyatomic gases present in the Earth’s 

troposphere, trap infrared radiation emitted by the Earth’s surface, leading to surface warming. 

This phenomenon is widely known as the “Greenhouse Effect” (Arrhenius, 1896), and the gases 

involved are known as “greenhouse gases” (GHGs). The term used for the role a particular 

atmospheric component, or any other component of the greater climate system, plays in altering 

the Earth’s radiative balance is “forcing.” In the past decade, the global atmospheric sciences 

and climate communities have made significant progress in determining the specific role that 

atmospheric O3 plays in forcing climate. 

The Intergovernmental Panel on Climate Change (IPCC) was founded in 1988 by the 

World Meteorological Society (WMO) and the United Nations Environmental Program (UNEP) 

to support the work of the Conference of Parties (COP) to the United Nations Framework 

Convention on Climate Change (UNFCCC). Drawing from the global climate and atmospheric 

sciences community for its authors and reviewers, the IPCC produces reports containing 

thorough assessments of the available peer-reviewed science regarding the physical climate 

system, past and present climate, and evidence of human-induced climate change. This section 

summarizes the reviews of the available information on the forcing properties of tropospheric O3 

as provided by the IPCC Third Assessment Report (IPCC, 2001a) and also describes some of the 

more recent developments on the subject. 

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The projected effects of global climate change will be briefly explained to provide the 

context within which O3 serves as a regional, and possibly global, anthropogenic pollutant. 

The concept of climate forcing is also explained, along with the factors that influence the extent 

of climate forcing by O3. The section concludes with a summary of the various estimates that 

have been placed on the amount of globally averaged forcing due to O3. 

10.3.1 The Projected Impacts of Global Climate Change 

The study of the atmospheric processes involved in global climate change, and its potential 

consequences for human health and global ecosystems, is an area of active research. The IPCC 

Third Assessment Report (TAR) is the most thorough evaluation currently available of the 

science concerning climate change. In addition to the first and second IPCC assessments in 

1990 and 1995, along with other IPCC reports, earlier assessments included those conducted by 

the UNEP (1986), the WMO (1988), the U.S. Environmental Protection Agency (1987), and 

others (e.g., Patz et al., 2000a,b). The reader is referred to those documents for a complete 

discussion of climate change science. An abbreviated list of the IPCC conclusions to date and a 

short discussion of the potential impacts of climate change on human health and welfare is 

provided here to serve as the context for the discussion of the role of the increasing tropospheric 

O3 concentration in climate change. 

According to various historic and modern measurement records, atmospheric GHG 

concentrations have increased dramatically in the past century and have been attributed to human 

activities. The IPCC TAR describes the scientific theory and evidence linking increases 

in GHGs to human activities (IPCC, 2001a). 

An increasing body of geophysical observations shows that the Earth is warming and that 

other climate changes are underway. These observations include the global surface temperature 

record assembled since the year 1860, the satellite temperature record begun in 1979, recorded 

changes in snow and ice cover since the 1950s, sea level measurements taken throughout the 

20th century, and sea surface temperature observations recorded since the 1950s. 

Observations (Levitus et al., 2005) show that ~84% of the total heating of the Earth System 

(oceans, atmosphere, continents, and cryosphere) over the last 40 years has gone into warming 

the oceans. Barnett et al. (2005) have reported the emergence of a clear pattern of ocean surface 

warming associated with anthropogenic GHGs. The authors constructed a model-based 

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fingerprint (i.e., a map of predicted changes in the vertical temperature profile of the Earth’s six 

major oceans) and compared this map to the newly upgraded and expanded ocean temperature 

data set (Levitus et al., 2005). They concluded that the warming signal far exceeds what would 

be expected from natural variability, a finding that was in compelling agreement with GHGforced model profiles. Other evidence of ocean warming includes a marked increase in the 

frequency, intensity, and persistence of the zonal atmospheric circulation shifts known as the 

El Niño-Southern Oscillation (ENSO) over the past 100 years. ENSO events occur when the 

tropical Pacific Ocean has accumulated a large, localized mass of warm water that interrupts 

cold surface currents along South America, altering precipitation and temperature patterns in the 

tropics, subtropics, and the midlatitudes. 

IPCC (1998, 2001a) reports also describe the results of general circulation model (GCM) 

studies predicting that human activities will alter the climate system in a manner likely to lead to 

marked global and regional changes in temperature, precipitation, and other climate properties. 

These changes are expected to increase the global mean sea level as well as increase the number 

of extreme weather events such as floods and droughts, increased wind speeds and precipitation 

intensity of tropical cyclones, and changes in soil moisture. These predicted changes can be 

expected to directly impact human health, ecosystems, and global economic sectors (e.g., 

hydrology and water resources, food and fiber production) (IPCC, 1998, 2001b). Table 10-1 

summarizes these projected impacts. 

Wide variations in the course and net impacts of climate change in different geographic 

areas are expected. In general, the projected changes constitute additional stressors on natural 

ecosystems and human societal systems already impacted by increasing resource demands, 

unsustainable resource management practices, and pollution. Some of the predicted changes 

include alterations in ecological balances; in the availability of adequate food, water, clean air; 

and in human health and safety. Poorer nations can be expected to suffer the most, given their 

limited adaptive capabilities. 

Although many regions are predicted to experience severe, possibly irreversible, adverse 

effects due to climate change, beneficial changes may also take place. For example, certain 

regions may benefit from warmer temperatures or increased CO2 fertilization, e.g., U.S. West 

Coast coniferous forests, and some Western rangelands. Specific benefits may include reduced 

energy costs for heating, reduced road salting and snow-clearance costs, longer open-water 

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Table 10-1. Examples of Impacts Resulting From Projected Changes in Extreme 

Climate Events 

Projected changes during the 21st Century 

in Extreme Climate Phenomena and their Representative Examples of Projected Impactsb 

Likelihooda (all high confidence of occurrence in some areasc

) 

Simple Extremes 

Higher maximum temperatures; more hot 

days and heat wavesd

 over nearly all land areas 

(very likelya

) 

Higher (increasing) minimum temperatures; 

fewer cold days, frost days, and cold wavesd 

over nearly all land areas (very likelya) 

More intense precipitation events 

(very likelya

 over many years) 

Complex Extremes 

Increased summer drying over most 

midlatitude continental interiors and 

associated risk of drought (likelya

) 

Increase in tropical cyclone peak wind 

intensities, mean and peak precipitation e intensities (likelya

 over some areas)

Intensified droughts and floods associated 

with El Niño events in many different regions 

(likelya

) (see also under droughts and intense 

precipitation events) 

Increased Asian summer monsoon precipitation 

variability (likelya

) 

Increased intensity of midlatitude storms 

(little agreement between current models)d 

C Increased incidence of death and serious illness in older age groups 

and urban poor 

C Increased heat stress in livestock and wildlife 

C Shift in tourist destinations 

C Increased risk of damage to a number of crops 

C Increased electric cooling demand and reduced energy supply reliability 

C Decreased cold-related human morbidity and mortality 

C Decreased risk of damage to a number of crops, and increased risk 

to others 

C Extended range and activity of some pest and disease vectors 

C Reduced heating energy demand 

C Increased flood, landslide, avalanche, and mudslide damage 

C Increased soil erosion 

C Increased flood runoff could increase recharge of some floodplain 

aquifers 

C Increased pressure on government and private flood insurance systems 

and disaster relief 

C Decreased crop yields 

C Increased damage to building foundations caused by ground shrinkage 

C Decreased water resource quantity and quality 

C Increased risk of forest fire 

C Increased risk to human life, risk of infections, disease epidemics, 

and many other risks 

C Increased coastal erosion and damage to coastal buildings and 

infrastructure 

C Increased damage to coastal ecosystems such as coral reefs and 

mangroves 

C Decreased agricultural and rangeland productivity in drought- and 

flood-prone regions 

C Decreased hydropower potential in drought-prone regions 

C Increased flood and drought magnitude and damages in temperate 

and tropical Asia 

C Increased risks to human life and health 

C Increased property and infrastructure losses 

C Increased damage to coastal ecosystems 

a

Likelihood refers to judgmental estimates of confidence used by TAR WGI: very likely (90-99% chance); likely

 (66-90% chance). Unless otherwise stated, information on climate phenomena is taken from the Summary for 

 Policymakers, TAR WGI. TAR WGI = Third Assessment Report of Working Group 1 (IPCC, 2001a). b

These impacts can be lessened by appropriate response measures. c

High confidence refers to probabilities between 67 and 95%.

d

Information from TAR WGI, Technical Summary. 

e

Changes in regional distribution of tropical cyclones are possible but have not been established. 

Source: IPCC (2001b). 

10-41 
seasons in northern channels and ports, and improved agricultural opportunities in the northern 

latitudes, as well as in the Western interior and coastal areas. For further details about the 

projected effects of climate change on a U.S.-regional scale, the reader is also referred to several 

regionally-focused reports (MARAT, 2000; Yarnal et al., 2000; NERAG, 2001; GLRAG, 2000), 

as well as a report on potential impacts of climate change on human health (Bernard et al., 

2001a,b). The IPCC report, “The Regional Impacts of Climate Change,” (IPCC, 1998) describes 

the projected effects of human-induced climate change on various regions of the globe, including 

Africa, the Arctic and Antarctic, the Middle East and arid Asia, Australasia, Europe, Latin 

America, North America, the small island nations, temperate Asia, and tropical Asia. 

While current climate models can successfully simulate the present-day annual mean 

global climate and the seasonal cycles on a continental scale, they have been less successful on a 

regional scale. Clouds and humidity, essential factors in defining local and regional (sub-grid 

scale) climate, are significantly uncertain (IPCC, 2001a). Due to modeling uncertainties, both in 

reproducing regional climate and in predicting the future economic activity that will govern 

future GHG emissions, the projected impacts discussed above are also uncertain. 

Findings from the U.S. Global Change Research Program (USGCRP) (NAST, 2000) and 

related reports illustrate the considerable uncertainties and difficulties in projecting likely 

climate change impacts at the regional or local scale. The USGCRP findings also reflect the 

mixed nature of projected potential climate change impacts, i.e., combinations of deleterious and 

beneficial effects, for U.S. regions and the variation of projected impacts across different 

regions. Difficulties in projecting region-specific climate change impacts are complicated by the 

need to evaluate the potential effects of regional- or local-scale changes in key air pollutants not 

only on global-scale temperature trends, but also on regional- or local-scale temperature and 

precipitation patterns. The EPA is currently leading a research effort that uses regional-scale 

climate models with the goal of identifying changes to O3 and PM concentrations that may occur 

in a warming climate. An assessment of the results of this effort will be available by the next 

review of the O3 NAAQS. This focused effort to determine the impact of a warming climate on 

criteria air pollution requires regional-scale models with improved skill in reproducing climate 

history and predicting change. Among the innovations being employed in this effort is the 

downscaling of global circulation model outputs to provide boundary conditions for model 

calculations at the regional scale (Liang et al., 2005). While focusing on projecting the impact of 

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a warming climate on regional O3 concentrations, the effort applied to improving regional-scale 

modeling will also lead to improved estimates of current and projected future impacts of 

tropospheric O3 on climate. 

10.3.2 Solar Energy Transformation and the Components of the Earth’s 

Climate System 

Mass, in any form, has the capacity to interact with solar and terrestrial radiation, but the 

manner in which it interacts with radiation is governed by its particular physical form and/or 

molecular properties. Water provides one of the most interesting examples of how physical form 

affects radiative properties. In its gaseous form, water is the most important GHG present in the 

climate system, because of its ability to absorb long-wave terrestrial radiation. Conversely, in its 

frozen form as snow or sea ice, water plays a very important role in the climate system by 

scattering UV and visible solar radiation back to space, i.e., decreasing the Earth’s net solar 

radiation receipts by increasing the Earth’s reflective properties (albedo). In its liquid aerosol 

form as clouds, water also greatly increases the Earth’s albedo. In its bulk liquid form as ocean 

water, it absorbs terrestrial radiation and represents the Earth’s most important reservoir of 

heat energy. 

The atomic composition and molecular structure of a gas determines the wavelengths of 

light it can absorb and, therefore, its role in defining the heat capacity of the atmosphere. Ozone 

and O2 provide examples of the relative importance of these molecular properties. While these 

molecules are both composed solely of oxygen atoms, their bond structures are distinct. Ozone 

has a three-atom, bent molecular structure, giving it the capacity to absorb terrestrial (infrared) 

wavelengths – making it a GHG. At any altitude, i.e., in the stratosphere or troposphere, O3 has 

the capacity to absorb UV radiation of 320 nm and shorter, further increasing the energyabsorbing capacity of the troposphere. Conversely, O2, due to its diatomic, linear structure, is 

limited to absorbing very short-wave UV light—and does so at altitudes too high to influence the 

climate system significantly. 

Each component of the climate system plays a role in absorbing, transforming, storing, 

dispersing, or scattering solar radiation. Weather is a tangible consequence of the transformation 

and dispersion of terrestrial radiation within the atmosphere. The term “weather” refers to the 

condition of the Earth’s atmosphere at a specific time and place. It is defined by several specific 

10-43
variables: the air temperature, air pressure, humidity, clouds, precipitation, visibility, and wind 

speed. The “climate” for a given place on the Earth’s surface is a long-term average of these 

variables accounting for daily and seasonal weather events. The frequency of extreme weather 

events is used to distinguish among climates that have similar averages (Ahrens, 1994). 

Climate components (including GHGs, land, oceans, sea ice, land ice and snow, 

atmospheric particles, vegetation, clouds, etc.) all contribute to the Earth’s heat capacity, i.e., its 

ability to absorb and retain solar energy. Changes in the properties (or mass) of these 

components will “force” the climate system in one direction or the other, i.e., warmer versus 

cooler. The transformation of atmospheric O2 into O3 by way of air pollution chemistry, 

enhances the heat capacity of the atmosphere. The principles behind the important concept of 

climate forcing are further described below. 

10.3.3 The Composition of the Atmosphere and the Earth’s 

Radiative Equilibrium 

The “greenhouse effect” is the term given to the decreased rate of reemission of absorbed 

solar energy due to the heat-retaining properties of the Earth’s atmosphere. According to simple 

radiative transfer theory, at thermal equilibrium, the Earth’s temperature should be near !15 °C. 

This is the temperature of a theoretical “black body” that is receiving and then reemitting 

342.5 Wm!2

, i.e., the globally averaged amount of full-spectrum solar energy absorbed and then 

reemitted by the Earth as infrared terrestrial radiation per square meter. In fact, satellite 

observations well above the atmosphere indicate that the Earth’s average planetary temperature 

is remarkably close to its theoretical black body value at !18 °C, a temperature at which liquid 

water ordinarily does not exist. 

At its surface, however, the Earth’s average temperature is +15 °C. The +33 °C 

temperature differential between the Earth’s planetary and surface temperatures is due to the 

presence of infrared (IR) radiation-absorbing components in the atmosphere such as water 

vapor, CO2, CH4, several other trace gases, and some types of particles and clouds. 

The atmosphere, when cloud-free, is largely transparent in the solar wavelength range. 

A small fraction of this radiation is absorbed and reemitted as black body radiation by dark 

atmospheric particles (IPCC, 2001a). However, the majority of clouds and particles, in part, 

10-44
offset the greenhouse effect by increasing the Earth’s albedo, thereby decreasing the overall 

amount of solar radiation absorbed by the Earth system. 

Ozone, SO2 and NO2 also absorb ultraviolet and near ultraviolet wavelengths, in addition to 

infrared radiation. Once absorbed by a gas molecule, the energy introduced by a photon may 

induce a photochemical reaction with the residual energy thermally exciting (heating) the 

products of the reaction. Alternatively, the energy introduced into the molecule by the photon 

may be dispersed amongst neighboring molecules via intermolecular collisions, or it may be 

reemitted in part as a lower energy (i.e., IR) photon. 

Radiation from the sun or the Earth’s surface that is absorbed by gases and particles is 

reemitted isotropically, i.e., it is equally likely to be emitted in all directions. Therefore, to a 

first approximation, half of the radiation trapped by the Earth’s atmosphere is reflected back to 

its surface. A portion of this radiation is transformed into the heat energy that drives the 

atmospheric processes that form the basis of weather and climate. Radiation that is not absorbed 

by gases and aerosols reaches the Earth’s surface where it is scattered (reflected) or absorbed, 

depending on the surface albedo. 

Successful modeling of the Earth’s climate and, therefore, the assessment of the extent of 

human-induced climate change and development of appropriate policy depend on the quality of 

available information on the relative efficiencies, amounts, and spatial and temporal distributions 

of the various radiatively active components of the atmosphere that absorb and/or reflect solar 

and terrestrial radiation, along with all the other nonatmospheric components of the Earth 

system. 

10.3.3.1 Forcing of the Earth’s Radiative Balance 

As mentioned earlier, the commonly used measure of the relative influence of a given 

component of the climate system on the Earth’s radiative balance is its radiative forcing (IPCC, 

2001a; Houghton et al., 1990). Radiative forcing, in Wm!2

, is a quantity that was developed by 

the climate modeling community as a first-order-only means of estimating relative effects of 

individual anthropogenic and natural processes on the energy balance within the climate system. 

When the effect of a particular component of the climate system is to reduce the amount of 

solar energy absorbed, usually by increasing the Earth’s albedo, this component is said to 

provide a “negative” forcing, measured in Wm!2

. The convention assigns a positive value to the 

10-45
forcing induced by climate system components that enhance the greenhouse effect or otherwise 

act to increase the heat-absorbing capacity of the Earth system. Purely reflective atmospheric 

aerosol, clouds, white rooftops, snow-covered land surfaces, and dense sea ice provide a 

negative forcing, while highly absorbing dark-colored atmospheric aerosols, GHGs, and 

increases in dark ocean surface area, due to the melting of sea ice sheets, positively force the 

climate system. 

Global and regional climate are roughly defined by the balance between the large number 

of positive and negative forcings induced by the many different components of the Earth system 

and any changes in the properties of those components due to natural processes or anthropogenic 

activities. Following a perturbation or added forcing, such as an increase in GHG concentrations 

or modification the Earth’s albedo through changes in land use, this balance is re-established via 

a complex redistribution of energy within the Earth system. Feedback mechanisms that are 

theorized but difficult to resolve at the quantitative level further complicate the prediction of the 

sensitivity of variables, such as surface temperature, to changes in forcing. 

A simple example of a positive feedback would be melting sea ice. As sea ice melts with 

increasing ocean temperatures, the dark ocean surface that is revealed is more efficient at 

absorbing IR radiation, further increasing the rate of warming. A negative feedback would be 

the formation of clouds over a moist, warming surface. As clouds form, less radiation is 

available to warm the surface, leading to cooling. The role of feedbacks in determining the 

sensitivity of climate to changes in radiative forcing is described in detail in the IPCC TAR 

(IPCC, 2001a). 

Discussions are presently underway within the climate community regarding a metric to 

replace forcing as the standard measure of climate impact—one that will account for more of the 

factors that determine the effectiveness of a specific change in altering climate. However, 

forcing remains the current standard (NRC, 2005). 

The IPCC has reported estimated values for forcing by individual radiatively active gases, 

and by particle-phase components of the atmosphere that were derived primarily through expert 

judgment incorporating the results of peer-reviewed modeling studies. The forcing estimates, 

shown in Figure 10-6, are global averages attributed to known GHGs, including O3, particles, 

anthropogenic cirrus clouds, land-use change, and natural solar flux variations. Uncertainty 

ranges are assigned to reflect the range of modeled values reported in those studies. The current 

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Figure 10-6. Estimated global mean radiative forcing exerted by gas and various particle 

phase species for the year 2000, relative to 1750. 

Source: IPCC (2001a). 

estimate of forcing due to long-lived, well-mixed, GHGs accumulated in the atmosphere from 

the preindustrial era (ca., 1750) through the year 2000 is +2.4 Wm!2

 ± 10% (IPCC, 2001a). 

An indication of the level of confidence in each of these estimates is given along the bottom of 

this figure, again reflecting the expert judgment of the IPCC. 

The IPCC reported a global average forcing value of 0.35 ± 0.15 Wm!2

 for tropospheric O3, 

based on model calculations constrained by climatological observations. The considerations and 

studies used to estimate this value are outlined below. Hansen and Sato (2001), accounting for 

uncertainties in pre-industrial emissions levels, more recently estimated a value of 0.5 ± 0.2 

Wm!2

 for forcing by O3. 

10.3.4 Factors Affecting the Magnitude of Climate Forcing by Ozone 

The radiative properties of O3 are distinct from those of other important GHGs in that it is 

capable of absorbing both UV and IR radiation. Furthermore, it is able to absorb long-wave 

radiation in a portion of the IR spectrum where water vapor does not absorb, i.e., the 9 to10 mm 

10-47
wavelength range, meaning that its ability to trap heat and force climate are unchanged by 

variations in humidity. Given its relatively short atmospheric lifetime in comparison to other 

GHGs, the distribution of tropospheric O3 is highly variable in geographic extent and time. 

These properties enhance the prospect of attributing a unique geographic and time-dependent 

pattern or fingerprint to forcing by O3. 

Due to human activities, tropospheric O3 is estimated to have provided the third largest 

increase in direct forcing since preindustrial times. It may also play a role in indirect forcing 

through its participation in the oxidative removal of other radiatively active trace gases, such as 

CH4 and the HCFCs. The direct and indirect forcing that tropospheric O3 imposes on the climate 

system depends upon its geospatial and temporal distribution, but it also depends upon its 

vertical position (altitude) in the atmosphere and the albedo of the underlying surface. 

10.3.4.1 The Global Burden of Tropospheric Ozone 

Little historical data exist that may be used to estimate the global O3 burden prior to 

industrialization, although a few late 19th-century measurements suggest that O3 has more than 

doubled in Europe during the 20th century. The insufficient data record on preindustrial 

tropospheric O3 distributions introduces a major uncertainty in the estimation of the change 

in O3-induced forcing since that period (IPCC, 2001a; Mickley et al., 2004a; Mickley et al., 

2001; Shindell and Faluvegi, 2002). 

Ozone reacts photochemically at time scales that are generally shorter than those for largescale mixing processes in the atmosphere. Concentrated O3 plumes evolve downwind of strong 

sources of its precursor pollutants: reactive nitrogen, CO, and non-methane hydrocarbons 

(NMHCs). The most important of these sources are midlatitude industrialized areas and tropical 

biomass burning. When viewed from above the atmosphere by satellite-borne spectrometers, O3 

enhancements appear as relatively localized air masses or regional-scale plumes, usually 

originating from industrialized areas or areas in which active biomass burning is underway. The 

IPCC (2001a) describes the efforts of several research teams who have analyzed data supplied by 

the satellite-borne Total Ozone Mapping Spectrometer (TOMS) and other remote-sensing 

instruments to map the global distribution of tropospheric O3 and to attempt to identify processes 

that influence the global tropospheric O3 budget (IPCC, 2001a). More recently, coincident 

observations of total O3 by TOMS and the Solar Backscattered UV (SBUV) instrument were 

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used by Fishman et al. (2003) to construct well-resolved spatial and temporal maps of the 

regional distribution of tropospheric O3. Their results were consistent with those reported by 

others, but with higher regional-scale resolution. They reported large O3 enhancements in the 

southern tropics in austral spring and in the northern temperate latitudes in the summer. The 

regional nature of high O3 concentrations was clearly visible in northeastern India, the eastern 

United States, eastern China, and west and southern Africa, each coincident with high population 

densities. Fishman et al. (2003) noted, as have the other groups cited above, significant 

interannual variability in the concentrations observed over these regions. In situ measurements 

of tropospheric O3 concentrations range from 10 ppb over remote oceans to 100 ppb in both the 

upper troposphere and in plumes downwind from polluted metropolitan regions (IPCC, 2001a). 

Ground-level concentrations in urban areas are often >100 ppb. In the Southern Hemisphere, 

one of the largest sources of O3 precursors is biomass burning. Biomass burning elevates O3 on 

large spatial scales, particularly in the tropical Atlantic west of the coast of Africa and in 

Indonesia. 

In its third assessment report, the IPPC estimates placed the global burden of tropospheric 

O3 at a highly uncertain 370 Tg, equivalent to an average column density of 34 Dobson Units (1 

DU = 2.687 × 1016 molecules/cm! 2) or a mean concentration of 50 ppb (IPCC, 2001a). 

Accounting for differences in levels of industrialization between the hemispheres, the average 

column burden in the Northern Hemisphere is estimated to be 36 DU, with the Southern 

Hemisphere estimated to average 32 DU. Due to its rapid photochemistry, individual surface 

measurements of tropospheric O3 cannot capture large-scale concentrations, nor will they 

represent the higher altitude concentrations. Dense surface and vertical measurements 

(ozonesondes) would be required to supplement available output from remote sensing 

instruments to provide the complete set of observations needed to derive a credible global O3 

budget. Such a measurement program appears, at present, to be impractical. 

10.3.4.2 Background Concentrations versus Regionally-Oriented Ozone Enhancements 

Vingarzan (2004) surveyed the air quality literature and reported that annual average 

background O3 concentrations at ground level in the Northern Hemisphere appear to range 

between 20 and 45 ppb, depending upon geographic location, elevation, and the influence of 

local sources. Fiore et al. (2003) modeled the U.S. continental O3 concentrations and found that 

10-49
surface background levels overlap the lower end of the range reported by Vingarzan (2004), e.g., 

15 to 35 ppb, with higher levels (40 to 50 ppb) arising at high-elevation sites due to the influence 

of the upper troposphere (See Chapter 3 and its associated annexes for a complete discussion of 

“policy relevant background [PRB]). Local- and regional-scale enhancements in O3 may be 

thought of as roughly superimposed upon these background levels, with the exception of longer 

stagnation events in which preexisting background O3 reacts away or is deposited as fresh O3 is 

produced from locally-emitted precursors. 

Lin et al. (2001) analyzed the EPA AIRS database for the 1980-1998 period and noted 

that O3 concentrations have declined at the high end of the probability distribution, consistent 

with the effects of emissions controls, but had increased at the low end of the distribution by 3 to 

5 ppb. They divided the monitoring data for the continental U.S. into 4 quadrants by geography 

and noted a pattern of increase for the Western states that might be attributed to the long-range 

transport of O3 precursors from Asia. They found, however, that the Northeastern quadrant had 

the highest increase in the low end of the concentration probability distribution, which could not 

be reasonably attributed to transboundary transport of O3 precursors. 

While not representing an ideal source of information for assessing the climatic effects 

of O3 within the continental United States, data from the large air-quality-focused ground-based 

monitoring network may be used to identify boundary-layer geospatial and temporal patterns 

in O3 concentrations for comparison to regional-scale chemistry/climate models. Extensive 

analysis of data available within the EPA AQS database can be found in Chapter 3 of this 

document, including an analysis showing the diurnal O3 concentration patterns for several large 

metropolitan areas with peak values ranging up to160 ppb (Los Angeles). Lehman et al. (2004) 

analyzed the AQS database of daily 8-h maximum O3 concentrations collected for 1,090 stations 

in the eastern half of the United States for the 1993 to 2002 period. They applied a rotated 

principle component analysis to a reasonably complete, spatially representative, nonurban subset 

of the database in order to identify coherent, regionally oriented patterns in O3 concentrations. 

Five spatially homogenous regions were identified: the U.S. Northeast, Great Lakes, MidAtlantic, Southwest (including Alabama, Louisiana, Texas, Oklahoma), and Florida. The 

Mid-Atlantic region displayed the highest mean concentration (52 ppb) of all of the regions 

analyzed, followed by the Great Lakes, Southwest, and Northeast regions with around 47 ppb. 

The average concentration derived for Florida was 41 ppb. The authors found strong 

10-50
correlations in measured concentrations among stations within the same region, suggesting that 

the geospatial patterns of pollutant emissions and meteorological activity may also have a 

regional orientation. These results suggest that these regions may define natural domains for 

regional scale modeling studies of the influence of O3 (as well as PM) on climate. 

10.3.4.3 Ozone Trends: Globally and in North America 

For the Northern Hemisphere, weekly continuous data are available from 1970 for only 

nine stations in the latitude range 36° N to 59° N (IPCC, 2001a). Available tropospheric O3 

measurements do not reveal a clear trend in concentration, while trends in the stratosphere are 

more readily identified. Different trends are seen at different locations for different periods, 

consistent with regional changes in pollutant emission, especially NOx. Logan et al. (1999) 

analyzed the composite record of mid-tropospheric O3 abundance from the nine-station network. 

A plot of data is shown in Figure 10-7. While no clear trend appeared for 1980 through 1996, 

the average level for second half of this record (about 57 ppb) is clearly greater than for the first 

half (about 53 ppb). The trend may be consistent with changes in regional NOx emission rates 

occurring due to pollution reduction efforts in developed countries being offset by increasing 

emissions in rapidly growing economies in Asia. The measurements shown in Figure 10-7 are 

concentrations observed at 4 to 7 km (mid-troposphere). Fewer locations have measured 

changes in the concentrations of O3 as a function of altitude. Fewer still are locations that have 

collected and maintained data records prior to 1970. The absence of historical data on the 

vertical distribution of O3 adds to the difficulty in estimating historical atmospheric burdens and 

trends in O3-related climate forcing. 

The IPCC (2001a) surveyed the results of published chemistry transport model (CTM) 

modelling studies (see Table 10-2) that estimated the global average increases in total column O3 

since the preindustrial era. Model estimates ranged from +7 to +12 DU. On the basis of these 

estimates, available measurements, and other analyses, the IPCC estimated that total column O3 

has increased by 9 DU, with a 67% confidence range of +6 to +13 DU. In some of the modelling 

studies, emissions scenarios predicted a further increase in column O3 due to growing emissions 

of O3 precursors. Fusco and Logan (2003) stated that, according to the models, increased NOx 

emissions from fossil fuel combustion have had the greatest effect on O3 in the lower 

troposphere since the 1970s. In addition, increases in background CH4 have also contributed as 

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Figure 10-7. Mid-tropospheric O3 abundance (ppb) in northern midlatitudes 

(36 °N-59 °N) for the years 1970 to 1996. Observations between 630 and 

400 hPa are averaged from nine ozonesonde stations (four in North America, 

three in Europe, two in Japan), following the data analysis of Logan et al. 

(1999). Values are derived from the residuals of the trend fit, with the trend 

added back to allow for discontinuities in the instruments. Monthly data 

(points) are shown with a smoothed 12-month mean (line). 

Source: IPCC (2001a). 

Table 10-2. CTM Studies Assessed by the IPCC for its Estimate of the Change in 

Global and Total Column O3 Since the Preindustrial Era 

Estimated Change in 

Column O3 in DU Model Used References 

7.9 GFDL Haywood et al. (1998) 

8.9 MOZART-1 Hauglustaine et al. (1998) 

8.4 NCAR/2D Kiehl et al. (1999) 

9.5 GFDL-scaled Levy et al. (1997) 

12 Harvard/GISS Mickley et al. (1999) 

7.2 ECHAM4 Roelofs et al. (1997) 

8.7 UKMO Stevenson et al. (2000) 

9.6 UIO Berntsen and Isaksen (1999) 

8 MOGUNTIA Van Dorland et al. (1997) 

Source: IPCC (2001a) 

10-52 
much as 20% to the increase in tropospheric O3 in the northern latitudes. Given its longer 

atmospheric residence time, CH4 can serve as an O3 precursor at much longer distances from its 

source than can other O3 precursors and, therefore, has a more uniform effect across the globe. 

Fusco and Logan (2003) found a 10% increase in O3 concentrations year-round over 

Canada, Europe, and Japan and a 20% increase for Japan and Europe during spring and summer. 

It was expected—but not found—that O3 concentrations over Japan would increase in line with 

emissions from China. The authors suggested that convective activity over Asia is stronger than 

that seen over other industrialized areas of the globe. Such a meteorological characteristic would 

result in an injection of pollutants into the free troposphere, allowing long-range transport to 

North America. Their suggestion is supported by evidence of increasing background 

concentrations within the United States (Fiore et al., 1998). 

NARSTO (2000), in its assessment of the available information on O3 pollution in North 

America, stated that no single pattern for trends in O3 over North America can be found in the 

available monitoring data. In the United States, the average 1-h concentration at surface 

monitoring sites decreased by 15% between 1986 and 1996, with most of the observed 

declines occurring in urban and urban-influenced locations. The largest declines occurred in 

Los Angeles, New York, and Chicago. Free tropospheric O3 concentrations appeared to hold 

steady, or only declined slightly, from the 1980s forward. 

In preparation for the IPCC TAR (IPCC, 2001a), research groups engaged in modeling 

global-scale tropospheric chemistry were invited to participate in a model intercomparison 

focusing on potential changes in the oxidative capacity of the atmosphere (OxComp), which 

included O3 concentrations for the 2000 to 2100 period. Participating groups employed the 

IPCC A2p scenario, i.e., including the highest emissions levels, to calculate the geospatial 

distribution of O3 up to 20 km. The predicted spatial distributions of O3 where quite variable, 

but the predictions for total column O3 density change fell within 9 DU of each other (11.4 to 

20.5 DU) in all cases and that was considered to be encouraging by the authors. Fusco and 

Logan (2003) pointed out that several unresolved issues may limit the ability of models in 

reproducing observed trends in tropospheric O3. Among these are the use of different 

meteorological inputs, photochemical reaction schemes, and predicted cloud cover—each 

contributing to different predictions in O3 production and loss rates. 

10-53
10.3.4.4 The Sensitivity of Ozone-Related Forcing Surface to Albedo 

The characteristics of the surface underlying an O3 enhancement play a role in the O3 

forcing effect. Highly reflective surfaces, such as light-colored deserts, sea ice and snow, scatter 

solar short wave (UV and visible) radiation. UV and visible radiation can then be absorbed, 

transformed into long-wave radiation, and reemitted in part back to the surface by tropospheric 

O3. Studies by two groups, Hauglustaine et al. (1998) and Mickley et al. (1999), have shown 

that industrial pollution that has been transported to the Arctic induces a high, regional O3­

related forcing due to the highly reflective underlying ice and snow surface. 

Liao et al. (2004) calculated that the maximum change in O3-related top-of-the-atmosphere 

forcing occurs over high albedo regions in high northern latitudes. Surface forcing was 

calculated to be greatest at high northern latitudes as well as at dust-source regions, which also 

tend to have high surface albedos. 

10.3.4.5 The Altitude Dependence of Forcing by Tropospheric Ozone 

Altitude plays an important role in the forcing effect of tropospheric O3 (IPCC, 1992; 

Gauss et al., 2003). The efficiency of IR absorption by O3 depends upon its temperature–at 

atmospheric temperatures that are low, relative to the Earth’s surface, it has the capacity to 

absorb more IR radiation than O3 at temperatures close to that of the surface. While this 

temperature effect applies to all GHGs, it introduces a complication for estimating forcing by O3, 

because O3 is not homogeneously mixed within the troposphere. Ozone forcing estimates must 

account for these difficult-to-predict vertical inhomogeneities. However, as part of the OxComp 

modeling intercomparison, Gauss et al. (2003) found that the overall forcing by O3 can be 

calculated within reasonable uncertainty simply on the basis of total column density. 

10.3.4.6 Co-occurrence of Ozone with Particulate Matter 

Analysis of the 2001 data from the AQS database showed infrequent co-occurrence of 

high PM2.5 and O3 concentrations (Chapter 3 of this document). For those cases when O3 

production is high, in combination with high PM concentrations, there is a suggestion in the 

literature that heterogeneous chemistry on PM surfaces may lead to reduced gas-phase O3. 

Liao et al. (2004) modeled heterogeneous chemistry taking place on PM, and found a significant 

10-54
titration of O3 and its NOx precursors. The importance of this titration effect remains an open 

question, given the difficulty in obtaining in situ measurements to validate model calculations. 

Liao et al. (2004) also estimated that forcing by BC, mineral dust, and organic carbon 

aerosols substantially offsets forcing by tropospheric O3, yielding an overall negative globally 

averaged forcing at both the top of the atmosphere and at the Earth’s surface. However, such 

estimates neglect the regional aspects of forcing by these individual pollutants. Elevated 

concentrations of these very different types of pollutants often appear independently of the 

others, such as with biomass burning plumes, Saharan dust, and organic aerosols associated 

with biogenic terpene emissions by forests. It is unlikely that a global average of the forcing 

effects of these individual pollutants will adequately capture their impacts on climate at the 

regional scale. 

10.3.5 Estimated Forcing by Tropospheric Ozone 

10.3.5.1 Direct Climate Forcing Due to Ozone 

The inhomogeneous distribution of O3 within the troposphere, coupled with the large 

uncertainty in the global O3 budget, significantly complicates the matter of estimating the global 

average direct forcing due to O3. The IPCC Third Assessment Risk (2001a) lists the results of 

several modeling studies that estimated the annual change in the relative forcing by O3 from 

preindustrial times. It was noted that the differences among the estimates were most likely due 

to differences in predicted O3 chemistry, including the emissions inventories used and the 

chemical process and transport mechanisms incorporated into the models, rather than by factors 

relating to radiative transfer. The IPCC intercomparison of the models and their results indicate 

that the uncertainties in estimated forcings due to O3 have decreased since the IPCC Second 

Assessment Report (1996). 

The O3-related forcings estimated by studies considered by the IPCC (2001a) are listed in 

Table 10-3. Ten of the listed estimates are based on global CTM calculations. One study was 

constrained by a climatology derived from observations. Given the differences in calculated 

total column O3 among the models, a normalized forcing (Wm!2

 per Dobson Unit of 

tropospheric O3 change) is listed in addition to the absolute forcing (Wm!2

) estimated by each 

model. Both clear sky (cloud-free) and total sky (including clouds) forcing estimates are listed. 

10-55
Table 10-3. Tropospheric O3 Change (O3) in Dobson Units (DU) Since Preindustrial 

Times, and the Accompanying Net (SW plus LW) Radiative Forcings (Wm!2

), 

After Accounting for Stratospheric Temperature Adjustment (using the Fixed 

Dynamical Heating Method). Estimates are Taken From the Published Literature. 

Normalized Forcings (norm.) Refer to Radiative Forcing per O3 Change (Wm!2

 per DU)

Estimated Global Average Forcing Due to Tropospheric Ozone 

Clear sky conditions Total sky conditions 

Reference O3 Net Net (norm.) Net Net (norm.) 

Berntsen et al. (1997) – [Reading model]

Stevenson et al. (1998)

Berntsen et al. (1997) – [Oslo model]

Haywood et al. (1998)

Kiehl et al. (1999)

Berntsen et al. (2000)

Brasseur et al. (1998)

Van Dorland et al. (1997)

Roelofs et al. (1997)

Lelieveld and Dentener (2000)

Hauglustaine et al. (1998)

Mean

7.600 0.310 0.041 0.280 0.037 

8.700 0.391 0.045 0.289 0.033 

7.600 0.390 0.051 0.310 0.041 

7.900 0.380 0.048 0.310 0.039 

8.400 0.379 0.045 0.320 0.038 

9.600 0.428 0.045 0.342 0.036 

— —– — 0.370 — 

8.070 0.443 0.055 0.380 0.047 

7.200 0.397 0.055 0.404 0.056 

— — — 0.420 — 

8.940 0.511 0.057 0.426 0.048 

8.224 0.403 0.049 0.343 0.042 

Source: IPCC (2001a). 

The largest O3-related forcings coincide with the strongest sources of tropospheric O3, 

which the models predict occur in the northern midlatitude regions (40° to 50° N) and reach as 

much as 1 Wm!2

 in the summer as well as in the tropics, and are related to biomass burning. 

In general, the estimates are comparable in magnitude and show similarity in geographic 

distribution. For total sky conditions, the range in globally and annually averaged 

tropospheric O3 forcing from all of these models is from 0.28 to 0.43 Wm!2

, while the 

normalized forcing is 0.033 to 0.056 Wm!2

 per DU. As expected, they are opposite in sign to 

the forcing estimated for sulfate aerosols, which scatter radiation. The range in normalized 

forcings emphasizes the differences in assumptions used by the different models. The 

10-56
tropospheric O3 forcing constrained by the observational climatology is 0.32 Wm!2

 for globally 

averaged, total sky conditions. As shown in Figure 10-6, the IPCC (2001a) concluded that 

0.35 ± 0.15 Wm!2 represents the most likely value for annually and globally-averaged forcing 

by tropospheric O3. Not included here is the study by Hansen and Sato (2001), that evaluated 

forcing by O3 with corrections made to the assumptions concerning pre-industrial O3 

concentrations and the effects of natural O3 precursors, especially NOx generated by lightning. 

Hansen and Sato (2001) concluded that a more likely range for globally averaged forcing by O3 

is 0.4 to 0.8 Wm!2, with 0.5 Wm!2 as their best estimate. Since the publication of the IPCC 

Third Assessment Report (2001a), new studies have been published that illuminate some of the 

regionally-relevant details associated with direct forcing by O3 (Mickley et al., 2004a; Liao et al., 

2004). Forcing by O3, due to its capacity for absorbing solar UV as well as solar and terrestrial 

IR radiation, can be divided into “shortwave” forcing and “long-wave” forcing. These forcings 

occur under different conditions. Shortwave forcing can only take place during daytime, while 

long-wave forcing can occur at all hours as a function of the diurnally varying concentration of 

atmospheric O3. As noted, earlier, unlike CO2, the absorption spectrum for O3 is distinct from 

that of water vapor—meaning that O3 will absorb and reemit long-wave radiation without 

interference by water under high humidity conditions. Mickley et al. (2004a) reported that, 

according to their modeling study, surface temperature response to the predicted O3 

enhancement since the preindustrial period differs greatly from that of the CO2 response, and 

that this difference can only be explained by the geographical distribution and absorption 

properties of O3. Liao et al. (2004) estimated globally averaged top-of-the-atmosphere separate 

short- and long-wave forcings to O3 to be 0.21 W/m2

 and 0.32 W/m2, respectively. 

10.3.5.2 Indirect Forcing Due to Ozone 

Ozone has an indirect climate forcing effect due to its role in the oxidative removal of 

other reactive GHGs, including CH4, hydrofluorocarbons (HFCs), and other reactive NMHCs. 

The primary actor in this effect is a second generation product of the photolysis of O3, the 

hydroxyl (OH) radical. Hydroxyl radicals are produced by way of a pair of reactions that start 

with the photodissociation of O3 by solar UV. 

10-57
(10-4) 

(10-5) 

Reactions with OH are the primary removal mechanism for CH4 and NMHCs as well as the 

pollutants NOx and CO. Methane and CO are in especially high abundance in the global 

atmosphere. OH is estimated to react with these two gases within 1 second of its formation. 

In addition to CH4, NOx, CO, and the NMHCs, OH concentrations are controlled by local 

concentrations of H2O (i.e., humidity) and the intensity of solar UV. Different atmospheric 

concentrations of the required precursors suggest that preindustrial OH concentrations were 

likely to have been different from present-day concentrations, but there is no consensus on the 

magnitude of this difference. Observations of global atmospheric concentrations of 

methylchloroform (CH3CCl3), a well-mixed tropospheric species that also reacts with OH, have 

been used to estimate OH abundances. Independent studies have shown overlapping trends for 

the period 1978 to 1994, but none of the trends are outside the given uncertainty ranges (0.5 ± 

0.6%/year) (Prinn et al., 1995; Krol et al., 1998). The IPCC (2001a) reported a range of +5% to 

!20% for predicted changes in global OH abundances. 

Given the difficulty in estimating global OH abundances in the past, present, and future, 

estimates of indirect forcing due to O3 have been difficult to obtain and are highly uncertain. 

10.3.5.3 Predictions for Future Climate Forcing by Anthropogenic Ozone 

Not surprisingly, CTM modeling attempts to predict future precursor emissions and 

resulting O3 abundances indicate that the largest future O3-related forcings will be related to 

population growth and economic development in Asia (Van Dorland et al., 1997; Brasseur et al., 

l998). The results of these modeling studies predict that the globally averaged total radiative 

forcing due to O3 from preindustrial times 0.66 Wm!2

 will rise to 0.63 Wm!2 by 2050. Chalita 

et al. (1996) predicted a change in the globally averaged radiative forcing from preindustrial 

times to 2050 of 0.43 Wm!2

. Stevenson et al. (l998) predicted an O3-related forcing of 0.48 

Wm!2

 in 2100. Applying the SRES scenario projecting the highest emissions out to the year 

2100 (IPCC, 2000), the OxComp model intercomparison study yielded a projected O3-induced 

forcing ranging from 0.40 to 0.78 Wm!2

. The authors concluded, given their prediction for 

10-58
forcing by well-mixed GHGs of 5.6 Wm!2

, that O3 would remain an important contributor to 

overall anthropogenic forcing well into the future. However, all of these predictions must be 

viewed with caution given the considerable uncertainties associated with the long-term economic 

activity projections required for such estimates. 

10.3.6 The Impact of a Warming Climate on Atmospheric Ozone 

Concentrations 

Evaluation of the potential impact of climate warming on U.S. air quality is currently 

underway. Initial modeling results reported by Mickley et al. (2004b) suggest that reduced 

cyclone frequency in a warmer climate will lead to increases in the severity of summertime 

pollution episodes. Cyclonic weather patterns are known to play an important role in ventilating 

pollution away from the surface. They note that compelling evidence is accumulating that the 

frequency of these cyclones has decreased over the past few decades. An early study by Jacob 

et al. (1993) found a correlation between O3 concentrations and temperature was due to the effect 

of temperature on atmospheric chemistry, biogenic emissions, and stagnation. 

10.3.7 Conclusion 

The general consensus within the atmospheric sciences community, as represented by the 

United Nations Intergovernmental Panel on Climate Change (IPCC), is that human activities 

have a discernable effect on the Earth’s climate. However, quantifying the extent of humaninduced forcing on climate requires detailed information about human-induced change on the 

components of the Earth System that govern climate. Tropospheric O3 is a well-known GHG, 

but information regarding its historical trends in concentration, its current and future 

atmospheric burden, and other critical details needed for estimating its direct and indirect 

forcing effects on the climate system are highly uncertain. 

The IPCC has estimated that the globally averaged forcing due to O3 is approximately 

0.35 ± 0.15 Wm!2

, with an updated value of 0.5 ± 0.2 Wm!2 provided by Hansen and Sato 

(2001). However, the most important role of O3 in climate is likely to be at the regional scale, 

adjacent to the sources of its chemical precursors. This expectation is consistent with satellite 

observations of high regional scale column O3 densities near large urban areas and large-scale 

biomass burning activity. Modeling studies evaluated by the IPCC have estimated that regional10-59
scale forcing due to O3 can approach 1 Wm!2

, or as much as 40% of the globally averaged 

forcing due to the well-mixed GHGs. While more certain estimates of the overall importance of 

global-scale forcing due to tropospheric O3 await further advances in monitoring and chemical 

transport modeling, the overall body of scientific evidence suggests that high concentrations 

of O3 on the regional scale could have a discernable influence on climate, leading to surface 

temperature and hydrological cycle changes. Confirming this effect requires improvement in 

regional-scale modeling—an activity that is currently underway. 

10-60
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11. OZONE EFFECTS ON MAN-MADE MATERIALS

Ozone (O3) and other photochemical oxidants react with many economically important 

man-made materials, decreasing their useful life and aesthetic appearance. Some materials 

known to be damaged by O3 include elastomers, fibers, dyes, and paints. This chapter provides a 

brief discussion of O3 effects on man-made materials, including denoting of damage mechanisms 

and, where possible, concentration-response relationships. Much of what is known about O3 

effects on man-made materials is derived from research conducted in the 1970's, 1980's, and 

early 1990's, with very little new research on the subject having been conducted since then. 

Since only very limited new information has been published on O3 effects on materials, this 

chapter mainly summarizes key information assessed in the previous 1996 Air Quality Criteria 

Document for Ozone and other photochemical oxidants (1996 O3 AQCD) (U.S. Environmental 

Protection Agency, 1996) and provides detailed discussion of the very limited new information 

that has become available since then. In the ensuing sections, discussion is focused on O3 effects 

on: elastomers (Sect 11.1); textiles and fabrics (11.2); dyes, pigments, and inks (11.3); artist’s 

pigments (11.4); and surface coatings (11.5). Evaluation of the relevance and economic 

importance of O3 materials damage information, as it affects productivity or cultural resources 

(such as museums), is beyond the scope of this chapter. The reader is referred to the previous 

criteria document (1996 O3 AQCD) for more detailed discussion of the earlier studies 

summarized below. 

11.1 ELASTOMERS 

The elastomeric compounds, natural rubber and synthetic polymers and copolymers of 

butadiene, isoprene, and styrene, are particularly susceptible to even low levels of O3. 

Elastomeric compounds are long chain unsaturated organic molecules. Ozone damages these 

compounds by breaking the molecular chain at the carbon-carbon double bond; a chain of three 

oxygen atoms is added directly across the double bond, forming a five-membered ring structure 

(Mueller and Stickney, 1970). The change in structure promotes the characteristic cracking of 

11-1
stressed/stretched rubber called “weathering.” A 5% tensile strain will produce cracks on the 

surface of the rubber that increase in number with increased stress/stretching. The rate of crack 

growth is dependent on the degree of stress, the type of rubber compound, concentration, time of 

exposure, velocity, and temperature (Bradley and Haagen-Smit, 1951; Lake and Mente, 1992) 

(Gent and McGrath, 1965). Once cracking occurs, there is further penetration, additional 

cracking, and eventually mechanical weakening or stress relaxation (U.S. Environmental 

Protection Agency, 1996). Razumovskii et al. (1988) demonstrated the effect of O3 on stress 

relaxation of polyisoprene vulcanizates. A decrease in stress (stress relaxation) is caused 

by O3-induced cracks in exposed elastomers resulting in irreversible changes in the elastomer 

dimensions and decreased tensile strength. 

To counteract O3 effects on elastomers, antiozonants and wax are often added to 

elastomeric formulations during processing. An antiozonant is an additive used to protect a 

polymer against the effects of O3-induced degradation and, hence, is used mainly in diene 

rubbers. Antiozonant protection works either (a) by providing a physical barrier to O3 

penetration via forming a thin surface film of an O3-resisting wax or (b) by chemically reacting 

with O3 or polymer ozonolysis products, as do aromatic diamines such as p-phenylene diamine 

derivatives. The antiozonant diffuses to the surface of the elastomeric material, where it reacts 

with O3 faster than O3 reacts to break the molecular chain and the carbon-carbon double bond, 

or the antiozonant diffuses to the surface of the material but is not reactive with O3 and serves as 

a protective coating against O3 attack. The antiozonant may also serve to scavenge O3 while also 

providing protective film against O3 attack (Andries et al., 1979; Lattimer et al., 1984). 

Most studies of O3 effects on elastomers were designed to evaluate the effectiveness of 

antiozonants in counteracting the rubber cracking produced by O3 exposure. Consequently, 

many of the studies were conducted using O3 concentrations notably higher than those 

typically found in the ambient air. Natural rubber strips exposed to high concentrations of O3 

(20,000 ppm) under stressed conditions cracked almost instantaneously and were broken within 

1 sec. When the O3 concentration was lowered (0.02 to 0.46 ppm), the time to required to 

produce cracks in the exposed rubber material was increased (Bradley and Haagen-Smit, 1951). 

Lake and Mente (1992) studied the effect of temperature on O3-induced elastomer cracking and 

antiozonant protection on natural rubber, epoxidised natural rubber, and two acrylonitrilebutadiene copolymers under constant strain. Temperatures ranged from -20 °C to +70 °C. 

11-2
The elastomers were exposed to 0.05 to 1,000 ppm O3 for 16 h. Ozone cracking decreased at 

lower ambient temperatures; however, diffusing of both chemical and wax antiozonants also 

slowed at the lower temperatures. Cracking was slightly increased at the higher temperatures, 

but the antiozonants offered more protection. 

Serrano et al. (1993) evaluated the appropriateness of using O3-induced elastomer cracking 

to estimate ambient O3 concentrations. Two vulcanized natural rubber compounds were exposed 

for 24 h to varying O3 concentrations under stressed conditions. Ozone concentrations were 

60, 80, 90, 100, and 120 ppb for durations of 2, 4, or 6 h. The 24 h average O3 concentrations 

ranged from 31 to 57.5 ppb. There was a clear relationship between the 24-h average O3 

concentration and the distribution of crack length frequencies on the rubber surface. Table 11-1 

gives the average 24-h O3 concentration and lengths for two vulcanized natural rubber strips. 

Table 11-1. Average 24-h Ozone Concentrations Producing the Highest Frequency of 

Cracks of a Certain Length in the Middle and Central Zones of the Rubber Test Strips 

1% Antiozonant 4010NA # 0.5% Antiozonant 4010NA 

Crack Length (mm) Middle Zones Central Zones Middle Zones Central Zones

 0.05 - 0.10 37.5 37.5 40.0 42.5

 0.10 - 0.15 45.0 48.0 48.0 53.0

 0.15 - 0.20 48.0 $57.5 $57.5 $57.5

 0.20 - 0.40 $57.5 $57.5 $57.5 $57.5 

Ozone concentrations given in ppb. 

Adapted from Serrano et al. (1993). 

11.2 TEXTILES AND FABRICS 

Ozone can damage textiles and fabrics by methods similar to those associated with 

elastomers. Generally, synthetic fibers are less affected by O3 than natural fibers; however, O3 

contribution to the degradation of textiles and fabrics is not considered significant (U.S. 

Environmental Protection Agency, 1996). A study reported by Bogaty et al. (1952) showed 

that O3 affects moistened cloth more than dry cloth. Scoured cotton duck cloth and 

11-3
commercially bleached cotton print cloth were exposed to 20 to 60 ppb O3 for 1,200 h (50 days). 

The rate of deterioration was measured by the changes in cuprammonium fluidity values and the 

fabric breaking strength. At the end of the 1,200-h exposure, there was a 20% loss in breaking 

strength. Table 11-2 list the changes in cuprammonium fluidity values for both fabrics. 

Table 11-2. Cuprammonium Fluidity of Moist Cotton Cloth Exposed 

to 20 to 60 ppb Ozone 

Duration of Exposure (h) Cuprammonium Fluidity (rhes) 

Duck Cloth 0 

200 

680 

960 

1200 

2.6 

2.8 

4.0 

6.8 

9.5 

Bleached Print Cloth 0 

200 

510 

650 

865 

1500 

8.2 

8.7 

9.4 

12.0 

12.7 

16.5 

Adapted from Bogaty et al. (1952). 

11.3 DYES, PIGMENTS, AND INKS 

Ozone fading of textile dyes is diffusion-controlled; the rate of fading is controlled by the 

diffusion of the dye to the fiber surface. Many textile dyes react with O3; however, the rate and 

severity of the O3 attack is influenced by the chemical nature of the textile fiber and the manner 

in which the dye is applied. Ozone molecules break the aromatic ring portion of the dye 

molecule, oxidizing the dye (U.S. Environmental Protection Agency, 1996). In case of aromatic 

azo dyes, O3 attacks the aromatic rings and electron rich nitrogen atoms (Matsui et al., 1988). 

Grosjean et al. (1987; 1988a,b) proposed a mechanism for reactions of O3 with indigo, 

thioindigo, and dibromoindigo, alazarin, and curcumin dyes under dark conditions. 

Ozone attaches to the dye molecule at the unsaturated carbon = carbon bond. An O3 adduct is 

formed (1,2,3-trioxolane), followed by scission of the carbon–carbon bond and the subsequent 

11-4
formation of the corresponding Criegee biradical. A similar mechanism was proposed for the 

reaction of O3 with triphenylmethane colorant Basic Violet 14. Ozone attacked Basic Violet 14 

at the carbon'carbon unsaturated bond and at the carbon&nitrogen unsaturated bond under dark 

conditions. Other members of the group of triphenylmethane colorants with unsaturated 

carbon–carbon bonds also are expected to be subject to O3 fading. Tripheylmethane colorants 

that are expected to be O3-fugitive include the amino-substituted cationic dyes (Malachite Green, 

Brilliant Green, Crystal Violet, Pararosaniline Chloride, Methyl Green, and others) (Grosjean 

et al., 1989). 

An indication that O3 caused textile dye fading was first reported by Salvin and Walker 

(1955). The researchers found that the fading was primarily the result of the destruction of the 

blue dye molecule. Drapes made of acetate, Arnel, and Dacron and dyed with anthraquinone 

blue dye exhibited a decrease in shade that was not accompanied by the characteristic reddening 

caused by NOx. Figures 11-1 and 11-2 demonstrate the effect of O3 exposure on nylon 6 yarn 

colored with several blue dyes. Nylon samples inside the home were located on a wall away 

from sunlight. Outside nylon samples were placed on a covered patio or under the eaves of the 

house to minimize exposure to sunlight and rain. Ozone concentrations ranged from 2 to 5 ppb 

outside and 0 to 2 ppb inside. The percent change in dye color was determined monthly by 

extraction and analysis of the remaining dye or by instrumental measurement of the color change 

(Haylock and Rush, 1978). 

11.4 ARTISTS’ PIGMENTS 

Several artists’ pigments are sensitive to fading and oxidation by O3 when exposed to 

concentrations found in urban areas (Shaver et al., 1983; Drisko et al., 1985; Whitmore et al., 

1987; Whitmore and Cass, 1988; Grosjean et al., 1993). The organic pigments that are O3 

fugitive include alizarin red pigments containing lakes of the polycyclic aromatic compound 

1,2-dihydroxyanthraquinone, blue-violet pigments containing substituted triphenylmethane 

lakes, indigo, and yellow coloring agents containing polyfunctional, polyunsaturated compounds 

such as curcumin (Grosjean et al., 1987). Because of the potential of O3 to damage works of art, 

recommended limits on O3 concentrations in museums, libraries, and archives are relatively low, 

ranging from 0.013 to 0.01 ppm. 

11-5
Figure 11-1. In-service fading of nylon 6 yarn inside house. ! = C.I. Disperse Blue 3; 

O = C.I. Basic Blue 22; > = C.I. Acid Blue 27; x = C.I. Disperse Blue 56;

Î = C.I. Acid Blue 232. 

Source: Haylock and Rush (1978). 

Experimental studies demonstrate a concentration × time (C × T) relationship between O3 

concentration, exposure time, and pigment fading. Cass et al. (1991) summarized some of the 

earlier research on the effects of O3 on artists’ pigments. In studies evaluating the effect of O3 on 

organic and inorganic watercolors and traditional organic pigments, only the traditional organic 

pigments showed measurable fading from O3 exposure. Of the inorganic pigments tested, 

only the arsenic sulfides showed O3-related changes. The pigments were exposed to 0.3 to 

0.4 ppm O3 for 3 mo in the absence of light, at 22 °C and 50% RH. The authors equated this 

exposure to a C × T of 6 to 8 years inside a Los Angeles museum with air conditioning but 

without a pollutant removal system. 

11-6
Figure 11-2. In-service fading of nylon 6 yarn outside house. ! = C.I. Disperse Blue 3; 

O = C.I. Basic Blue 22; > = C.I. Acid Blue 27; x = C.I. Disperse Blue 56;

Î = C.I. Acid Blue 232. 

Source: Haylock and Rush (1978). 

Whitmore and Cass (1988) studied the effect of O3 on traditional Japanese colorants. Most 

of these compounds are insoluble metal salts that are stable in light and air. Suspensions or 

solutions of the colorants were airbrushed on hot-pressed watercolor paper or silk cloths. 

A sample of Japanese woodblock print also was included in the analysis. Samples were exposed 

to 0.4 ppm O3 at 22 °C, 50% relative humidity, in the absence of light for 12 wk. Changes in 

reflectance spectra were used to evaluate the effect of O3 exposure on colorant fading. Among 

the colorants tested on paper, curmin, indigo, madder lake, and lac lake were the most sensitive 

to O3 exposure. Gamboge was relatively insensitive to O3. The blue and green areas of the 

sample from the woodblock print was very reactive due to the indigo dye O3 sensitivity. The 

11-7
other colorants, red, yellow, and purple, showed very little sensitivity to O3. The textiles dyes 

that reacted with O3 were indigo, alone or in combination with several yellow dyes. 

Ye et al. (2000) reported the rate of O3 fading of traditional Chinese plant dyes. Twelve 

different colorants were applied to watercolor paper and silk and exposed to 0.4 ppm O3 at 

25 °C, at 50% RH, in the absence of light for 22 wks. Dye fading was greater when the colorant 

was applied to the watercolor paper compared to the silk cloth due to the darker initial depth of 

the shade, the greater saturation of the colorant throughout the cloth. Tumeric, gromwell, and 

violet on paper was particularly reactive. Tangerine peel was moderately reactive and sappan 

wood, dalbergia wood, Chinese gall, indigo, and Chinese yellow cork tree were slightly reactive 

to O3. Black tea was not reactive to O3. The colorants on silk samples showing color changes 

were gromwell, sappan wood, gardenia, tummeric, and violet. Figures 11-3 and 11-4 

demonstrate the color change of the various colorants on watercolor paper and silk. 

Artists’ pigments also have exhibited fading when exposed to a mixture of photochemical 

oxidants. Grosjean et al. (1993) exposed 35 artists’ pigments to a mixture of photochemical 

oxidants consisting of O3, nitrogen dioxide (NO2), and peroxyacetyl nitrate (PAN) for 12 wks. 

Weekly average photochemical concentrations were 200 ppb for O3, 56 ± 12 to 99 ± 24 for NO2, 

and 11 ± 3 to 18 ± 2 for PAN. All exposures were carried out at room temperature in the 

absence of light. To determine the effect of humidity on pigment fading, the relative humidity 

was increased from 46% after 8 weeks of exposure to 83% for a 2 week period and then returned 

to 46% for the remainder of the exposure. 

Table 11-3 lists the artists’ pigment and degree of fading. Eleven of the pigments tested 

exhibited negligible color change, 12 had small color changes, 3 had modest color changes, and 

9 exhibited substantial color changes. Fading of Disperse Blue 3 and Reactive Blue 2 were 

likely the result of NO2 exposure, and the fading of triphenylmethanes is consistent with 

exposure to nitric acid formed under high humidity conditions. Fading of the indigos was 

dominated by O3 exposure and curcumin was faded by all of the photochemicals studied. 

Increasing the relative humidity resulted in a substantial color change for all of the pigments, 

with the exception of curcumin and indigo. 

11-8
Figure 11-3. Observed color changes for natural colorant-on-paper systems during 

exposure to 0.40 ppm O3 at 25 °C ± 1 °C, 50% RH, in the absence of light. 

Source: Ye et al. (2000). 

11-9 
Figure 11-4. Observed color changes for natural colorant-on-site during exposure to 

0.40 ppm O3 at 25 °C ± 1 °C, 50% RH, in the absence of light. 

Source: Ye et al. (2000). 

11-10 
Table 11-3. Color Change After 12 Weeks of Exposure to a Mixture of 

Photochemical Oxidants 

Colorant* Color Change ()E units)† 

Chemical Functionality or 

Chemical Composition 

Acid Red 37 (17045)‡ 11.7 ± 0.5 Aminophenyl-substituted azo dye, 

sulfonate salt 

Acid Yellow 65 ‡ 1.8 ± 0.5 Nitro- and phenyl-substituted azo dye, 

sulfonate salt 

Alizarin Carmine 1.8 ± 0.2 Alizarin lake 

Alizarin Crimson 1.4 ± 0.2 Alizarin lake 

(Pigment Red 83) 

Aurora Yellow (77199) 0.5 ± 0.1 Cadmium sulfide 

Basic Fuschin (42510)‡ 33.4 ± 3.0 Amino-substituted triphenylmethane 

Brilliant Green (42040)‡ 20.6 ± 2.1 Amino-substituted triphenylmethane 

Brown Madder 1.7 ± 0.1 Alizarin lake 

Cadmium Yellow (77199) 0.4 ± 0.1 Cadmium sulfide 

Carmine 1.8 ± 0.2 Lake of cochineal (substituted 

anthraquinone) 

Chrome Yellow (77600)‡ 1.7 ± 1.2 Lead chromate 

Copper phthalocyamne 1.0 ± 0.1 Copper phthalocyanine 

(Pigment Blue 15) 

Crimson Lake 3.5 ± 0.3 Alizarin lake 

Curcumin (Natural Yellow 3) 15.2 ± 2.6 1,7 bis (4-hydroxy-3-methoxyphenyl)-

1,6-heptadiene-3,5-dione 

Disperse Blue 3 10.8 ± 0.1 Amino-substituted anthraquinone 

French Ultramarine Blue 0.8 ± 0.3 

Gamboge (Natural Yellow 24) 0.4 ± 0.1 Gambogic acid 

Hooker's Green Light 1.5 ± 0.4 Chlorinated copper phthalocyanine plus 

ferrous beta naphthol derivative 

Indigo (a formulation) 1.1 ± 0.1 Alizarin lake plus lampblack plus copper 

plthalocyanine 

Indigo carmine ‡ 14.0 ± 1.9 5,5-indigo disulfonic acid, sodium salt 

Indigo (73000)‡ 64.1 ± 4.5 

Mauve 3.6 ± 0.5 Lake of triphenyl methane (basic fuschin) 

plus copper phthalyocyanine 

New Gamboge 0.9 ± 0.1 Arylamide yellow (CI 11680) plus 

toluidine red 

11-11
Table 11-3 (cont’d). Color Change After 12 Weeks of Exposure to a 

Mixture of Photochemical Oxidants 

Colorant* Color Change ()E units)† 

Chemical Functionality or 

Chemical Composition 

Pararosaniline base (42500)‡ 25.6 ± 4.7 Amino-substituted triphenylmethane 

Payne's Grey 1.0 ± 0.1 Alizarin lake plus prussian blue plus 

lampblack plus ultramarine blue 

Permanent Magenta 1.1 ± 0.1 Quinacridone 

Permanent Rose 2.0 ± 0.1 Quinacridone 

Prussian Blue 0.7 ± 0.2 

1.6 ± 0.3 

Ferric ferrocyanide 

Prussian Green 0.9 ± 0.2 Arylamide yellow plus prussian blue 

Purple Lake 2.3 ± 0.3 Alizarin lake 

Reactive Blue 2 (61211)‡ 14.4 ± 1.1 Amino-substituted anthraquinone, 

sulfonate salt 

Rose Carthane (12467) 0.8 ± 0.2 Arylamide (Pigment Red 10) plus 

xanthene (Pigment Red 90) 

Rose Doré 2.0 ± 0.2 Quinacridone plus Yellow 3 

Thioindigo Violet (73312)‡ 1.9 ± 1.2 Chlorinated thioindigo 

Winsor Yellow (11680) 0.5 ± 0.2 Arylamide yellow 

* On watercolor paper unless otherwise indicated. Color Index (CI) names or CI numbers are given in

parentheses.

† Mean ± one standard deviation for triplicate samples calculated from the parameters L*, a*, and b* measured

with the color analyzer.

‡ On Whatman 41 paper. 

Source: Grosjean et al. (1993). 

11.5 SURFACE COATINGS 

Ozone will act to erode some surface coatings (paints, varnishes, and lacquers). However, 

many of the available studies on O3 degradation of surface coatings do not separate the effects 

of O3 from those of other pollutants or environmental factors such as weather, humidity, and 

temperature. Campbell et al. (1974) attempted to demonstrate an O3-related effect on oil house 

paint, acrylic latex coating, alkyd industrial maintenance coating, urea alkyd coil coating, 

and nitrocellulose/acrylic automotive paint. Painted test panels were exposed to 100 and 

11-12
1,000 ppb O3 in a xenon arc accelerated weathering chamber for up to 1,000 h. Using weight 

loss as a measure of O3-induced erosion, the researchers concluded that all of the paints tested 

suffered degradation in the presence of O3 and that the automotive finish suffered the most O3­

induced degradation. When O3 degradation was measured using scanning electron microscopy, 

the oil house paint and latex coating samples showed erosion above that seen with clean air but 

only at the highest exposure level. No effects were noted for the automotive paint. The other 

painted surfaces were not evaluated. 

Spence et al. (1975) studied the effect of air pollutants and relative humidity on oil based 

house paint, acrylic latex house paint, acrylic coil coating, and vinyl coil coating under 

laboratory conditions. Test panels were exposed in weathering chambers equipped with a xenon 

are light for simulating sunlight to low and high levels of O3 (0.08 and 0.5 ppm), sulfur dioxide 

(0.03 and 0.5 ppm), and nitrogen dioxide (0.05 and 0.5 ppm) and relative humidity (50 and 

90%). Samples were exposed for a total of 1000 h. The exposure cycle consisted of 20 min of 

dew and 20 min of light. The effects of the exposure on the painted surfaces were measured by 

weight loss and loss in film thickness. The acrylic coil coating had the lowest erosion rate of the 

surface coatings tested. However, O3 was the only pollutant that had a significant effect on the 

surface erosion. Sulfur dioxide and relative humidity were significant factors in the erosion of 

oil base house paints and vinyl coil coating. The findings for acrylic latex house paint were not 

reported. 

11.6 CONCLUSIONS 

Ozone and other photochemical oxidants react with many economically important 

man-made materials, decreasing their useful life and aesthetic appearance. Some materials 

known to be damaged by O3 include elastomers, fibers and dyes, and paints. Most studies have 

been on single compounds rather than complex materials. 

The elastomeric compounds, natural rubber and synthetic polymers and copolymers of 

butadiene, isoprene, and styrene, are particularly susceptible to even low concentrations of O3. 

Ozone damages these compounds by breaking the molecular chain at the carbon-carbon double 

bond; a chain of three oxygen atoms is added directly across the double bond. The change in 

structure promotes the characteristic cracking of stressed/stretched rubber called “weathering.” 

11-13
Tensile strain produces cracks on the surface of the rubber that increase in number with 

increased stress/stretching. 

The rate of crack growth is dependent on the degree of stress, the type of rubber 

compound, O3 concentration, time of exposure, O3 velocity, and temperature. After initial 

cracking, there is further O3 penetration, resulting in additional cracking and, eventually, 

mechanical weakening or stress relaxation. 

Ozone can damage textiles and fabrics by methods similar to those associated with 

elastomers. Generally, synthetic fibers are less affected by O3 than natural fibers; however, O3 

contribution to the degradation of textiles and fabrics is not considered significant . 

Ozone fading of textile dyes is a diffusion-controlled process; the rate of fading is 

controlled by the diffusion of the dye to the fiber surface. Many textile dyes react with O3. The 

rate and severity of the O3 attack is influenced by the chemical nature of the textile fiber and the 

manner in which the dye is applied. 

Several artists’ pigments are also sensitive to fading and oxidation by O3 when exposed to 

concentrations found in urban areas. 

11-14
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11-15
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11-16 
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