Policy Assessment for the Review of the 

Ozone National Ambient Air Quality 

Standards 
 

 

DISCLAIMER

This document has been reviewed by the Office of Air Quality Planning and Standards 

(OAQPS), U.S. Environmental Protection Agency (EPA), and approved for publication. This 

OAQPS Policy Assessment contains conclusions of the staff of the OAQPS and does not 

necessarily reflect the views of the Agency. Mention of trade names or commercial products is 

not intended to constitute endorsement or recommendation for use. 
EPA-452/R-14-006 

August 2014 

Policy Assessment for the Review of the Ozone National Ambient Air 

Quality Standards 

U.S. Environmental Protection Agency 

Office of Air and Radiation 

Office of Air Quality Planning and Standards 

Health and Environmental Impacts Division 

Ambient Standards Group 

Research Triangle Park, North Carolina 27711 
This page left intentionally blank. 
i 

TABLE OF CONTENTS 

LISTS OF FIGURES..............................................................................v 

LIST OF TABLES..................................................................................ix 

LIST OF ACRONYMS AND ABBREVIATIONS...........................................xi 

EXECUTIVE SUMMARY....................................................................ES-1 

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

1.1 PURPOSE ........................................................................................................ 1-1 

1.2 BACKGROUND ............................................................................................. 1-3 

1.2.1 Legislative Requirements ......................................................................... 1-3 

1.2.2 History of O3 NAAQS Reviews ............................................................... 1-5 

1.2.3 Current O3 NAAQS Review .................................................................. 1-10 

1.3 GENERAL APPROACH FOR REVIEW OF THE STANDARDS ............. 1-12 

1.3.1 Approach for the Primary Standard ....................................................... 1-13 

1.3.1.1 Approach Used in the Last Review ............................................... 1-14 

1.3.1.2 Approach for the Current Review ................................................. 1-17 

 Consideration of the Scientific Evidence ...................... 1-20 

 Consideration of Exposure and Risk Estimates ............ 1-24 

 Considerations Regarding Ambient O3 Concentration 

Estimates Attributable to Background Sources ............. 1-26 

1.3.2 Approach for the Secondary Standard ................................................... 1-27 

1.3.2.1 Approach Used in the Last Review ............................................... 1-28 

1.3.2.2 Approach for the Current Review ................................................. 1-33 

 Consideration of the Scientific Evidence ...................................... 1-36 

 Consideration of Exposure and Risk Estimates and Air Quality 

 Analyses ...................................................................................... 1-39 

 1.3.2.2.1 Considerations Regarding Ambient O3 Concentration Estimates 

 Attributable to Background Sources...................................1-41 

1.3.3 Organization of this Document .............................................................. 1-42 

1.4 REFERENCES .............................................................................................. 1-43 

2 O3 MONITORING AND AIR QUALITY ............................................................. 2-1

2.1 O3 MONITORING ............................................................................................... 2-1

2.1.1 O3 Monitoring Network .............................................................................. 2-1

2.1.2 Recent O3 Monitoring Data and Trends ...................................................... 2-3

2.2 EMISSIONS AND ATMOSPHERIC CHEMISTRY .......................................... 2-9

2.3 AIR QUALITY CONCENTRATIONS ............................................................. 2-10

2.4 BACKGROUND O3 .......................................................................................... 2-12

2.4.1 Seasonal Mean Background O3 in the U.S. .............................................. 2-17

2.4.2 Seasonal Mean Background O3 in the U.S. as a Proportion of Total O3 .. 2-19

2.4.3 Daily Distributions of Background O3 within the Seasonal Mean ........... 2-20
ii 

2.4.4 Proportion of Background O3 in 12 Urban Case Study Areas .................. 2-24

2.4.5 Influence of Background O3 on W126 levels ........................................... 2-25

2.4.6 Estimated Magnitude of Individual Components of Background O3 ....... 2-27

2.4.7 Summary ................................................................................................... 2-30

2.5 REFERENCES ................................................................................................... 2-32 

3     ADEQUACY OF THE CURRENT PRIMARY STANDARD.......................3-1

3.1 EVIDENCE-BASED CONSIDERATIONS ........................................................ 3-2

3.1.1 Modes of Action .......................................................................................... 3-2

3.1.2 Nature of Effects ......................................................................................... 3-7

3.1.2.1 Respiratory Effects – Short-term Exposures ...................................... 3-8

3.1.2.2 Respiratory Effects – Long-term Exposures .................................... 3-36

3.1.2.3 Total Mortality – Short-term Exposures .......................................... 3-45

3.1.2.4 Cardiovascular effects – Short-term Exposure ................................ 3-49

3.1.3 Adversity of Effects .................................................................................. 3-52

3.1.4 Ozone Concentrations Associated With Health Effects ........................... 3-56

3.1.4.1 Concentrations in Controlled Human Exposure Studies and in 

 Epidemiologic Panel Studies..............................................3-56

3.1.4.2 Concentrations in Epidemiologic Studies – Short-term Metrics ..... 3-60

3.1.4.3 Concentrations in Epidemiologic Studies – “Long-term” Metrics .. 3-74

3.1.5 Public Health Implications ........................................................................ 3-77

3.1.5.1 At-Risk Populations ......................................................................... 3-77

3.1.5.2 Size of At-Risk Populations and Lifestages in the United States .... 3-87

3.1.5.3 Averting Behavior ............................................................................ 3-89

3.2 AIR QUALITY-, EXPOSURE-, AND RISK-BASED CONSIDERATIONS .. 3-90

 3.2.1 Consideration of the Adjusted Air Quality Used in Exposure and Risk 

 Assessments .............................................................................................. 3-90

3.2.2 Exposure-Based Considerations ............................................................... 3-93

3.2.3 Risk-Based Considerations ..................................................................... 3-103

3.2.3.1 Risk of Lung Function Decrements ............................................... 3-103

3.2.3.2 Estimated Health Risks Associated with Short- or Long-Term O3 

 Exposures, Based on Epidemiologic Studies.........................3-113

 3.3 CASAC ADVICE AND PUBLIC COMMENTERS’ VIEWS ON THE 

 ADEQUACY OF THE CURRENT STANDARD ......................................... 3-124

3.4 STAFF CONCLUSIONS ON ADEQUACY OF PRIMARY STANDARD .. 3-128

3.5 REFERENCES ................................................................................................ 3-137

4 CONSIDERATION OF ALTERNATIVE PRIMARY STANDARDS ............... 4-1

4.1 INDICATOR .................................................................................................... 4-1

4.2 AVERAGING TIME ....................................................................................... 4-2

4.3 FORM .............................................................................................................. 4-5

4.4 LEVEL ............................................................................................................. 4-8

4.4.1 Evidence-based Considerations ................................................................ 4-9

4.4.2 Air Quality-, Exposure-, and Risk-Based Considerations ...................... 4-21

4.4.2.1 Exposure-Based Considerations .................................................... 4-21

4.4.2.2 Risk-Based Considerations: Lung Function .................................. 4-29
iii 

4.4.2.3 Risk-Based Considerations: Epidemiology-Based Mortality and 

Morbidity ....................................................................................... 4-36

4.5 CASAC ADVICE AND PUBLIC COMMENTERS’ VIEWS ON 

ALTERNATIVE STANDARDS ................................................................... 4-46

4.6 STAFF CONCLUSIONS ON ALTERNATIVE PRIMARY STANDARDS 

FOR CONSIDERATION .............................................................................. 4-48

4.7 KEY UNCERTAINTIES AND AREAS FOR FUTURE RESEARCH AND 

DATA COLLECTION .................................................................................. 4-70

4.8 SUMMARY OF STAFF CONCLUSIONS ON PRIMARY STANDARD .. 4-73

4.9 REFERENCES .............................................................................................. 4-76 

5 ADEQUACY OF THE CURRENT SECONDARY STANDARD.......................5-1

5.1 NATURE OF EFFECTS AND BIOLOGICALLY RELEVANT EXPOSURE 

METRIC...........................................................................................................5-1

5.2 FOREST TREE GROWTH, PRODUCTIVITY AND CARBON 

STORAGE.....................................................................................................5-11

5.2.1 Evidence-based Considerations..............................................................5-12

5.2.2 Exposure/Risk-based Considerations......................................................5-29

5.3 CROP YIELD LOSS.....................................................................................5-43

5.3.1 Evidence-based Considerations..............................................................5-43

5.3.2 Exposure/Risk-based Considerations......................................................5-48

5.4 VISIBLE FOLIAR INJURY..........................................................................5-51

5.4.1 Evidence-based Considerations...............................................................5-53

5.4.2 Exposure- and Risk-based Considerations...............................................5-60

5.5 OTHER WELFARE EFFECTS.....................................................................5-69

5.5.1 Forest Susceptibility to Insect Infestation................................................5-69

5.5.2 Fire Regulation.........................................................................................5-70

5.5.3 Ozone Effects on Climate........................................................................5-71

5.5.4 Additional Effects....................................................................................5-72

5.6 CASAC ADVICE..........................................................................................5-73

5.7 STAFF CONCLUSIONS ON ADEQUACY OF SECONDARY 

STANDARD..................................................................................................5-75

5.8 REFERENCES...............................................................................................5-89

6 CONSIDERATION OF ALTERNATIVE SECONDARY STANDARDS ......... 6-1

6.1 INDICATOR .................................................................................................... 6-1

6.2 FORM AND AVERAGING TIME ................................................................. 6-2

6.3 LEVEL ............................................................................................................. 6-8

6.4 CONSIDERATION OF PROTECTIVENESS OF REVISED PRIMARY 

STANDARD ................................................................................................. 6-37

6.5 CASAC ADVICE .......................................................................................... 6-40

6.6 STAFF CONCLUSIONS ON ALTERNATIVE STANDARDS .................. 6-44

6.7 SUMMARY OF CONCLUSIONS ON THE SECONDARY STANDARD 6-57

6.8 KEY UNCERTAINTIES AND AREAS FOR FUTURE RESEARCH AND 

DATA COLLECTION .................................................................................. 6-58

6.9 REFERENCES .............................................................................................. 6-62
iv 

APPENDICES

Appendix 2A. Supplemental Air Quality Modeling Analyses of Background O3 .................. 2A-1 

Appendix 2B. Monitoring Data Analysis of Relationships Between Current Standard and W126 

Metric......................................................................................2B-1 

Appendix 2C. Inter-annual Variability in W126 Index Values: Comparing Annual and 3-Year 

Average Metrics (2008-2010) ............................................................................2C-1 

Appendix 3A. Recent Studies of Respiratory-Related Emergency Department Visits and Hospital 

Admissions...............................................................................3A-1 

Appendix 3B. Ambient O3 Concentrations in Locations of Health Studies ..............................3B-1 

Appendix 5A. O3-Sensitive Plant Species Used by Some Tribes ............................................. 5A-1 

Appendix 5B. Class I Areas Below Current Standard And Above 15 ppm-hrs..................5B-1 

Appendix 5C. Expanded Evaluation of Relative Biomass and Yield Loss........................5C-1 

 
v 

List of Figures 

Figure 1-1.     Overview of approach to reviewing the primary standard. .................................. 1-19

Figure 1-2.     Overview of approach to reviewing the secondary standard. .............................. 1-35

Figure 2-1. Map of U.S. ambient O3 monitoring sites reporting data to EPA during the 2006-

2010 period. ........................................................................................................... 2-3

Figure 2-2. Trend in U.S. annual 4th highest daily maximum 8-hour O3 concentrations in ppb, 

2000 to 2012. Solid center line represents the median value across monitoring sites, 

dashed lines represent 25th and 75th percentile values, and top/bottom lines 

represent 10th and 90th percentile values. ............................................................. 2-4

Figure 2-3.   Map of 8-hour O3 design values in ppb for the 2009-2011 period. ....................... 2-5

Figure 2-4.   Map of 8-hour O3 design values in ppb for the 2010-2012 period. ....................... 2-5

Figure 2-5. Trend in U.S. annual W126 concentrations in ppm-hrs, 2000 to 2012. Solid center 

line represents the median value across monitoring sites, dashed lines represent 

25th and 75th percentile values, and top/bottom lines represent 10th and 90th 

percentile values. .................................................................................................... 2-6

Figure 2-6.   Map of 2009-2011 average annual W126 values in ppm-hrs. ............................... 2-7

Figure 2-7.   Map of 2010-2012 average annual W126 values in ppm-hrs. ............................... 2-7

Figure 2-8. Trend in the May to September mean of the daily maximum 8-hour ozone 

concentrations before (dotted red line) and after (solid blue line) adjusting for yearto-year variability in meteorology. ........................................................................ 2-8

Figure 2-9. Map of 2007 CMAQ-estimated seasonal mean natural background O3 levels (ppb) 

from zero-out modeling. ...................................................................................... 2-18

Figure 2-10. Map of 2007 CMAQ-estimated seasonal mean North American background O3

levels (ppb) from zero-out modeling. .................................................................. 2-19

Figure 2-11. Map of 2007 CMAQ-estimated seasonal mean United States background O3 levels 

(ppb) from zero-out modeling. ............................................................................. 2-19

Figure 2-12. Map of site-specific ratios of U.S. background to total seasonal mean O3 based on 

2007 CMAQ zero-out modeling. ......................................................................... 2-22

Figure 2-13. Map of site-specific ratios of apportionment-based U.S. background to seasonal 

mean O3 based on 2007 CAMx source apportionment modeling. ....................... 2-22

Figure 2-14. Distributions of absolute estimates of apportionment-based U.S. Background (all 

site-days), binned by modeled MDA8 from the 2007 source apportionment 

simulation. ............................................................................................................ 2-23

Figure 2-15. Distributions of the relative proportion of apportionment-based U.S. Background to 

total O3 (all site-days), binned by modeled MDA8 from the 2007 source 

apportionment simulation. ................................................................................... 2-23

Figure 2-16. Fractional influence of background sources to W126 levels in four sample 

locations. Model estimates based on 2007 CMAQ zero-out modeling. .............. 2-27

Figure 2-17. Differences in seasonal mean O3 (ppb) between the NAB and NB scenarios. 2-28

Figure 2-18. Percent contribution of U.S. anthropogenic emissions to total seasonal mean 

MDA8 O3 levels, based on 2007 source apportionment modeling. ................... 2-30

Figure 3-1. Modes of action/possible pathways underlying the health effects resulting from 

inhalation exposure to O3. (Adapted from U.S. EPA, 2013, Figure 5-8) .............. 3-4
vi 

Figure 3-2. Percent increase in respiratory-related hospital admission and emergency 

department visits in studies that presented all-year and/or seasonal results. ....... 3-34

Figure 3-3. Summary of mortality risk estimates for short-term O3 and all-cause 

(nonaccidental) mortality. .................................................................................... 3-47

Figure 3-4.     Concentration-response function for asthma hospital admissions over the 

distribution of area-wide averaged O3 concentrations (adapted from Silverman and 

Ito, 2010). ............................................................................................................ 3-66

Figure 3-5. Concentration-response function for pediatric asthma emergency department visits 

over the distribution of averaged, population-weighted 8-hour O3 concentrations 

(reprinted from Strickland et al., 2010). .............................................................. 3-68

Figure 3-6. Exposure-Response relationship between risk of death from respiratory causes and 

ambient O3 concentration study metric (Jerrett et al., 2009). .............................. 3-76

Figure 3-7. Percent of children estimated to experience one or more exposures of concern at or 

above 60, 70, 80 ppb with air quality adjusted to just meet the current standard - 

Averaged Over 2006 to 2010. ............................................................................. 3-96

Figure 3-8. Percent of children estimated to experience one or more exposures of concern at or 

above 60, 70, 80 ppb with air quality adjusted to just meet the current standard - 

Worst-Case Year from 2006 to 2010. .................................................................. 3-97

Figure 3-9. Percent of children estimated to experience two or more exposures of concern at or 

above 60, 70, 80 ppb with air quality adjusted to just meet the current standard - 

Averaged Over 2006 to 2010. .............................................................................. 3-98

Figure 3-10. Percent of children estimated to experience two or more exposures of concern at or 

above 60, 70, 80 ppb with air quality adjusted to just meet the current standard - 

Worst-Case Year from 2006 to 2010. .................................................................. 3-99

Figure 3-11. Percent of school-aged children (5-18 yrs) estimated to experience one or more 

days with FEV1 decrements > 10, 15, or 20% with air quality adjusted to just meet 

the current standard – Averaged over 2006 to 2010 .......................................... 3-107

Figure 3-12. Percent of school-aged children (5-18 yrs) estimated to experience one or more 

days with FEV1 decrements > 10, 15, or 20% with air quality adjusted to just meet 

the current standard – Worst-Case Year from 2006 to 2010 ............................. 3-108

Figure 3-13. Percent of school-aged children (aged 5-18 yrs) estimated to experience two or 

more days with FEV1 decrements > 10, 15, or 20% with air quality adjusted to just 

meet the current standard – Averaged over 2006 to 2010 ................................. 3-109

Figure 3-14. Percent of school-aged children (5-18 yrs) estimated to experience two or more 

days with FEV1 decrements > 10, 15, or 20% with air quality adjusted to just meet 

the current standard - Worst-Case Year from 2006 to 2010 .............................. 3-110

Figure 3-15. Percent of all-cause mortality associated with O3 for air quality adjusted to just 

meet the current standard. .................................................................................. 3-115

Figure 3-16. Estimated O3-associated deaths attributable to various area-wide average O3

concentrations, with air quality adjusted to just meet current standard. ............ 3-116

Figure 3-17.   Percent of baseline respiratory mortality estimated to be associated with long-term 

O3. .................................................................................................................... 3-118

Figure 4-1. Percent of children estimated to experience one or more exposures of concern at or 

above 60, 70, or 80 ppb for air quality adjusted to just meet the current and 

potential alternative standards (averaged over 2006 to 2010) ............................. 4-23
vii 

Figure 4-2. Percent of children estimated to experience one or more exposures of concern at or 

above 60, 70, or 80 ppb for air quality adjusted to just meet the current and 

potential alternative standards (worst-case year from 2006 to 2010) ................. 4-24

Figure 4-3. Percent of children estimated to experience two or more exposures of concern at or 

above 60, 70, or 80 ppb for air quality adjusted to just meet the current and 

potential alternative standards (averaged over 2006 to 2010) ............................. 4-25

Figure 4-4. Percent of children estimated to experience two or more exposures of concern at or 

above 60, 70, or 80 ppb for air quality adjusted to just meet the current and 

potential alternative standards (worst-case year from 2006 to 2010) ................. 4-26

Figure 4-5. Percent of children estimated to experience one or more O3-induced lung function 

decrements greater than 10, 15, or 20% for air quality adjusted to just meet the 

current and potential alternative standards (averaged over 2006 to 2010) .......... 4-30

Figure 4-6. Percent of children estimated to experience one or more O3-induced lung function 

decrements greater than 10, 15, or 20% for air quality adjusted to just meet the 

current and potential alternative standards (worst-case year from 2006 to 2010)

 ......................................................................................................................... ...4-31

Figure 4-7. Percent of children estimated to experience two or more O3-induced lung function 

decrements greater than 10, 15, or 20% for air quality adjusted to just meet the 

current and potential alternative standards (averaged over 2006 to 2010) .......... 4-32

Figure 4-8. Percent of children estimated to experience two or more O3-induced lung function 

decrements greater than 10, 15, or 20% for air quality adjusted to just meet the 

current and potential alternative standards (worst-case year from 2006 to 2010)

 ......................................................................................................................... ...4-33

Figure 4-9. Estimates of Total Mortality Associated with Short-Term O3 Concentrations in 

Urban Case Study Areas (Air Quality Adjusted to Current and Potential alternative 

standard levels) – Total Risk ............................................................................... 4-38

Figure 4-10.   Estimates of O3-Associated Deaths Attributable to Full Distribution of 8-Hour 

Area-Wide O3 Concentrations and to Concentrations at or above 20, 40, or 60 ppb 

- Deaths Summed Across Urban Case Study Areas ............................................ 4-40

Figure 4-11. Estimates of Respiratory Hospital Admissions Associated with Short-Term O3

Concentrations in Urban Case Study Areas (Air Quality Adjusted to Current and 

Potential alternative standard levels) – Total Risk .............................................. 4-41

Figure 4-12. Estimates of Respiratory Mortality Associated with long-term O3 Concentrations 

in Urban Case Study Areas (Air Quality Adjusted to Current and Potential 

alternative standard levels) – Total Risk ............................................................. 4-43

Figure 4-13. Estimates of O3-Associated Deaths Attributable to Full Distributions of 8-Hour 

Area-Wide O3 Concentrations and to Concentrations at or above 20, 40, or 60 ppb 

O3 - Deaths Summed Across Urban Case Study Areas and Expressed Relative to a 

Standard with a Level of 75 ppb ......................................................................... 4-51 

Figure 5-1. A) Relative biomass loss in seedlings for 12 studied species using composite 

functions in response to seasonal O3 concentrations in terms of seasonal W126 

index values, Y-axis scale for RBL values represents 0% up to 100% (U.S. EPA 

2014, Figure 6-2)..................................................................................................5-25 

Figure 5-1. B) Expanded view of relative biomass loss in seedlings for 12 studied species 

using composite functions in response to seasonal O3 concentrations in terms of 
viii 

lower range of seasonal W126 index values, Y-axis scale for RBL values 

represents 0% up to 10% (U.S. EPA 2014, Figure 6-2).......................................5-26 

Figure 5-2. Relationship of tree seedling percent biomass loss with seasonal W126 index. 

(From U.S. EPA 2014, Figure 6-5)......................................................................5-33 

Figure 5-3. Relative biomass loss of Ponderosa Pine for air quality adjusted to just meet the 

current standard (U.S. EPA 2014, Figure 6-8).....................................................5-34 

Figure 5-4. Relative yield loss in crops using the composite functions for 10 studied species in 

response to seasonal O3 concentrations in terms of seasonal W126 index values, Yaxis scale for RYL values represents 0% up to 100% (U.S. EPA 2014, Figure 6-

3)..........................................................................................................................5-45 

Figure 5-5. Cumulative proportion of biosites with any foliar injury present, by moisture 

category (U.S. EPA 2014, Figure 7-10)...............................................................5-64 
ix 

List of Tables 

Table 1-1. Summary of primary and secondary O3 NAAQS promulgated during the period from 

1971 to 2008. ............................................................................................................. 1-5

Table 2-1 Comparison of the two model methodologies used to characterize background ozone 

levels. ....................................................................................................................... 2-15

Table 2-2. Seasonal mean MDA8 O3 (ppb), seasonal mean apportionment-based USB 

contribution (ppb), and fractional apportionment-based USB contribution to total O3

(all site-days) in the 12 REA urban case study areas (%). ....................................... 2-25

Table 2-3. Seasonal mean MDA8 O3 (ppb), seasonal mean apportionment-based USB 

contribution (ppb), and fractional apportionment-based USB contribution to total O3

(site-days > 60 ppb) in the 12 REA urban study areas (%). .................................... 2-25

Table 2-4. Fractional contribution of apportionment-based USB in the 12 REA urban study areas 

(%), using the means and medians of daily MDA8 fractions (instead of fractions of 

seasonal means). ...................................................................................................... 2-25

Table 2-5. Seasonal mean MDA8 O3 (ppb), seasonal mean USB (ppb), and USB/Total ratio (all 

site-days) in the 12 REA urban case study areas (%). ............................................. 2-25

Table 3-1. Group mean results of controlled human exposure studies that have evaluated 

exposures to ozone concentrations below 75 ppb in young, healthy adults. ........... 3-58

Table 3-2. Panel studies of lung function decrements with analyses restricted to O3

concentrations below 75 ppb. .................................................................................. 3-60

Table 3-3. U.S. and Canadian epidemiologic studies reporting O3 health effect associations in 

locations that would have met the current standard during study periods. ............. 3-63

Table 3-4. Distributions of daily 8-hour maximum ozone concentrations from highest monitors 

over range of 2-day moving averages from composite monitors (for study area 

evaluated by Silverman and Ito, 2010) .................................................................... 3-67

Table 3-5. Distribution of daily 8-hour maximum ozone concentrations from highest monitors 

over range of 3-day moving averages of population-weighted concentrations (for 

study area evaluated by Strickland et al., 2010) ...................................................... 3-69

Table 3-6. Number of study cities with 4th highest daily maximum 8-hour concentrations greater 

than 75 ppb, for various cut-point analyses presented in Bell et al. (2006) ............ 3-74

Table 3-7. Prevalence of asthma by age in the U.S. ................................................................. 3-88

Table 4-1 Numbers of epidemiologic study locations likely to have met potential alternative 

standards with levels of 70, 65, and 60 ppb ............................................................ 4-14

Table 4-2 Number of study cities with 3-year averages of 4th highest 8-hour daily max 

concentrations greater than 70, 65, or 60 ppb, for various cut-point analyses presented 

in Bell et al. (2006) .................................................................................................. 4-16

Table 4-3 Seasonal averages of 1-hour daily max O3 concentrations in U.S. urban case study 

areas for recent air quality and air quality adjusted to just meet the current and 

potential alternative standards. ................................................................................ 4-20

Table 4-4 Summary of Estimated Exposures of Concern for Potential Alternative O3 Standard 

Levels of 70, 65, and 60 ppb in Urban Case Study Areas ....................................... 4-49

Table 4-5 Summary of Estimated Lung Function Decrements for Potential Alternative O3

Standard Levels of 70, 65, and 60 ppb in Urban Case Study Areas ....................... 4-50 
x 

Table 5-1. O3-Sensitive Trees, Their Uses and Relative Sensitivity.........................................5-20

Table 5-2. O3 concentrations in Class I areas during period from 1998 to 2012 that met the 

current standard and where three-year average W126 index value was at or above 15 

ppm-hrs...................................................................................................................5-28 

Table 5-3. Exposure, risk and ecosystem services analyses related to tree growth, productivity 

and carbon storage..................................................................................................5-30 

Table 5-4. Summary of methodology by which national surface of 3-year average W126 index 

values was derived for each air quality scenario.....................................................5-31 

Table 5-5. Number of Counties with Tree Species Exceeding 2% Relative Biomass Loss.....5-37 

Table 5-6. Exposure, risk and ecosystem services analyses related to crop yield.....................5-48 

Table 5-7. Visible foliar injury incidence in four National Wildlife Refuges...........................5-57 

Table 5-8. Exposure, risk and ecosystem services analyses related to visible foliar injury......5-61 

Table 5-9. Benchmark criteria for O3 exposure and relative soil moisture used in screening-level 

assessment of parks (from U.S. EPA 2014, Table 7-6)..........................................5-65 

Table 6-1. Tree seedling biomass loss and crop yield loss estimated for O3 exposure over a 

season......................................................................................................................6-11 

Table 6-2. Percent of assessed geographic area exceeding 2% weighted relative biomass loss in 

WREA air quality scenarios....................................................................................6-20 

Table 6-3. Number of Class I areas (of 145 assessed) with weighted relative biomass loss 

greater than 2%.......................................................................................................6-22 

Table 6-4. Estimated mean yield loss (and range across states) due to O3 exposure for two 

important crops.......................................................................................................6-23 

Table 6-5. Estimated effect of O3-sensitive tree growth-related impacts on the ecosystem 

services of air pollutant removal and carbon sequestration in five urban case study 

areas........................................................................................................................6-25 
xi 

List of Acronyms and Abbreviations

Act Clean Air Act 

ACS American Cancer Society 

AHR Airway hyperresponsiveness 

ANF Atrial natriuretic factor 

AOT40 Seasonal sum of the difference between an hourly concentration at the 

threshold value of 40 ppb, minus the threshold value of 40 ppb 

AOT60 Seasonal sum of the difference between an hourly concentration at the 

threshold value of 60 ppb, minus the threshold value of 60 ppb 

APEX Air Pollutants Exposure model 

APHEA Air Pollution and Health: A European Approach 

APHENA Air Pollution and Health: A European and North American Approach 

AQCD Air Quality Criteria Document 

AQI Air Quality Index 

AQRV Air quality related value 

AQS Air Quality System 

ARG Arginase 

AspenFACE Aspen Free Air gas Concentration Enrichment Facility 

ATS American Thoracic Society 

BALF Bronchoalveolar Lavage Fluid 

BI Biosite Index 

BRFSS Behavioral Risk Factor Surveillance System 

C Concentration 

CAA Clean Air Act 

CAL FIRE California Department of Forestry and Fire Protection 

CAMx Comprehensive Air Quality Model with Extensions 

CAMP Childhood Asthma Management Program 

CAR Centriacinar region 

CASAC Clean Air Scientific Advisory Committee 

CASTNET Clean Air Status and Trends Network 

CAT Catalase 

CBSA Core-based statistical area 

CD Criteria Document 

CDC Centers for Disease Control 

CFR Code of Federal Regulations 

CH4 Methane 

CHD Coronary Heart Disease 

CHF Congestive Heart Failure 

CHS Children’s Health Study 

CI Confidence interval 

CMAQ Community Multi-scale Air Quality Model 

CO Carbon monoxide 

CO2 Carbon dioxide 

COPD Chronic obstructive pulmonary disease 

C-R Concentration-response 
xii 

CSA Combined Statistical Area 

CSTR Continuous stirred tank reactor 

CVD Cardiovascular disease 

DHEW Department of Health, Education, and Welfare 

ED Emergency department 

ELF Extracellular Lining Fluid 

EPA Environmental Protection Agency 

E-R Exposure-response 

eVNA Enhanced Voronoi Neighbor Averaging

FACE Free-air CO2 (and ozone) enrichment system 

FASOMGHG Forest and Agricultural Optimization Model – Greenhouse gas version 

FEM Federal Equivalent Method 

FeNO Exhaled nitric oxide fraction 

FEV1 Forced Expiratory Volume for 1 second 

FHM/FIA Forest Health Monitoring /Forest Inventory and Analysis Program 

FHWAR National Survey of Fishing, Hunting, and Wildlife-Associated Recreation 

FIA USDA Forest Inventory and Analysis Program 

FLAG Federal Land Managers’ Air Quality Related Values Workgroup 

FR Federal Register 

FRM Federal Reference Method 

FVC Forced Vital Capacity 

GEOS Goddard Earth Observing System 

GIS Geographic Information Systems 

GRSM Great Smoky Mountains National Park 

GSTM1 Glutathione-S-transferase polymorphism M1 genotypes 

GSTP1 Glutathione-S-transferase polymorphism P1 genotypes 

HA Hospital Admission 

HDDM Higher Order Direct Decoupled Method 

HEI Health Effects Institute 

HMOX1 Heme oxygenase-1 polymorphism 

HO Heme oxygenase 

HR Heart rate 

HREA Health Risk and Exposure Assessment 

HRV Heart rate variability 

ICD-9 International Classification of Disease - 9th revision 

ICU Intensive care unit 

IgE Immunoglobulin E 

IL Interleukin 

I/R Ischemia-reperfusion 

ISA Integrated Science Assessment 

Max maximum 

MDA8 Maximum daily 8-hour ozone average 

MMTCO2e Million metric tonnes of carbon dioxide equivalents 

MPO Myeloperoxidase 

MSA Metropolitan Statistical Area 

NAAQS National ambient air quality standards 
xiii 

NAB North American background 

NB Natural background 

NCDC National Climatic Data Center 

NCLAN National Crop Loss Assessment Network 

NCORE sites National Core multi-pollutant monitoring sites

NHLBI National Heart, Lung, and Blood Institute 

NMMAPS National Morbidity, Mortality, and Air Pollution Study 

NO Nitric oxide 

NO2 Nitrogen Dioxide 

NOx Nitrogen Oxides 

NOAA National Oceanic and Atmospheric Administration 

NQO1 NAD(P)H-quinone oxidoreductase genotype 

NRC National Research Council 

NRCS Natural Resources Conservation Service 

NSRE National Survey on Recreation and the Environment 

NWR National wildlife refuges 

O3 Ozone 

OC Organic carbon 

OIF Outdoor Industry Foundation 

OIRA Office of Information and Regulatory Affairs 

OMB Office of Management and Budget 

OR Odds ratio 

OTC Open-top chamber 

PA Policy Assessment 

PAMS Photochemical Assessment Monitoring Stations 

PAN Peroxyacetyl nitrate 

PAPA Public Health and Air Pollution in Asia 

PDSI Palmer Drought Severity Index 

PEFR Peak Expiratory Flow Rate 

ppm Parts per million 

ppm-hrs part per million-hours 

ppb Parts per billion 

PM Particulate matter 

PM2.5 Particles generally less than or equal to 2.5 μm in diameter 

PM10 Particles generally less than or equal to 10 micrometers (μm) in diameter 

PMN Polymorphonuclear leukocyte 

POMS Portable Ozone Monitoring System 

PRB Policy relevant background 

QA Quality assurance 

RBL Relative biomass loss 

ROMO Rocky Mountain National Park 

RR Relative risk 

SEKI Sequoia and Kings Canyon National Parks 

SES Socioeconomic status 

SIP State Implementation Plans 

SLAMS State and Local Monitoring Stations 
xiv 

SO2 Sulfur Dioxide 

SOx Sulfur Oxides 

SoyFACE Soybean Free Air gas Concentration Enrichment Facility 

SP Staff Paper 

SPMS Special Purpose Monitoring Stations 

STE Stratospheric-tropospheric exchange 

SUM00 Season sum of all hourly average concentrations 

SUM06 Seasonal sum of all hourly average concentrations ≥ 0.06 ppm 

TLC Total Lung Capacity 

TNF Tumor Necrosis Factor 

UNEP United Nations Environmental Programme 

U.S. United States 

USB United States background 

UV Ultraviolet 

 Ventilation rate 

VNA Voronoi neighbor Averaging 

VOC Volatile Organic Compounds 

W126 Cumulative integrated exposure index with a sigmoidal weighting function 

WHO World Health Organization 

wRBL Weighted relative biomass loss 

WREA Welfare Risk and Exposure Assessment 

WTP Willingness to pay 
 ES-1

EXECUTIVE SUMMARY 

This Policy Assessment (PA) has been prepared by staff in the Environmental Protection 

Agency’s (EPA) Office of Air Quality Planning and Standards (OAQPS) as part of the Agency’s 

review of the primary (health-based) and secondary (welfare-based) national ambient air quality 

standards (NAAQS) for ozone (O3). The current O3 standards were established in 2008 at the end 

of the previous review cycle. These standards include a primary O3 standard of 75 ppb,1

 and a 

secondary O3 standard set identical to the primary standard. These 2008 standards are now under 

review, as required by sections 108 and 109 of the Clean Air Act (Act). The PA presents 

analyses and staff conclusions regarding the policy implications of the key scientific and 

technical information that informs this review. Staff conclusions are presented regarding the 

adequacy of the current standards and potential alternative standards appropriate for 

consideration. Staff analyses in this PA are based on the scientific and technical information, 

including the uncertainties and limitations related to this information, assessed and presented in 

the Integrated Science Assessment for Ozone (ISA), the Health Risk and Exposure Assessment 

for Ozone (HREA), and the Welfare Risk and Exposure Assessment for Ozone (WREA). The 

PA is intended to “bridge the gap” between the relevant scientific evidence and technical 

information and the judgments required of the EPA Administrator in determining whether to 

retain or revise the current standards. Development of the PA is also intended to facilitate advice 

and recommendations on the standards to the Administrator from an independent scientific 

review committee, the Clean Air Scientific Advisory Committee (CASAC), as provided for in 

the Act. Staff considerations and conclusions in this final PA have been informed by comments 

and recommendations from CASAC, and by public comments. 

Health Effects and Review of the Primary Standard 

A longstanding and comprehensive evidence base, stronger now than in the last review, 

documents the effects of O3 exposures on human health. It is well-understood that secondary 

oxidation products, which develop as a result of O3 exposure, initiate numerous responses at the 

cellular, tissue, and whole organ level of the respiratory system. These key initiating events have 

the potential to result in a variety of adverse respiratory effects, as well as effects outside the 

respiratory system (e.g., cardiovascular effects). Ozone inhalation poses the greatest risk to 

people in certain lifestages (i.e., children, older adults), people with asthma, people with certain 

genetic variants (related to oxidative stress and inflammation), people with diets limited in 

 

1

 The level of the O3 standard is specified as 0.075 ppm rather than 75 ppb. However, in this PA we refer to ppb, 

which is most often used in the scientific literature and in the ISA, in order to avoid the confusion that could result 

from switching units when discussing the evidence in relation to the standard level. 
 ES-2

certain nutrients (antioxidant vitamins C and E), and people experiencing the largest exposures 

(e.g., outdoor workers, children). The evidence from animal toxicology and controlled human 

exposure studies indicates that higher exposure concentrations and repeated exposures lead to a 

greater prevalence of effects and increasingly severe effects, including increased susceptibility to 

other respiratory stressors, among exposed populations, especially these at-risk populations. 

As an initial matter in this PA, staff concludes that reducing ambient O3 concentrations to 

meet the current standard of 75 ppb will provide important improvements in public health 

protection. This initial conclusion is based on (1) the strong body of scientific evidence 

indicating a wide range of adverse health outcomes attributable to exposures to O3

concentrations found in the ambient air and (2) estimates indicating decreased O3 exposures and 

health risks upon meeting the current standard, compared to recent air quality. 

Strong support for this initial conclusion is provided by controlled human exposure 

studies of respiratory effects, and by quantitative estimates of exposures of concern and lung 

function decrements based on information in these studies. Analyses in the HREA estimate that 

the percentages of children (i.e., all children and children with asthma) in urban case study areas2

experiencing exposures of concern, or experiencing abnormal and potentially adverse lung 

function decrements, are consistently lower for air quality that just meets the current O3 standard 

than for recent air quality. The HREA estimates such reductions consistently across the urban 

case study areas and across years evaluated, and throughout various portions of individual urban 

case study areas, including in urban cores and in the portions of case study areas surrounding 

urban cores. These reductions in exposures of concern and O3-induced lung function decrements 

reflect consistent reductions in relatively high O3 concentrations (i.e., those in the upper portions 

of the distribution of ambient concentrations) following reductions in precursor emissions to 

meet the current standard. Thus, populations in both urban and non-urban areas would be 

expected to experience important reductions in O3 exposures and O3-induced lung function risks 

upon meeting the current standard. 

Support for this initial conclusion is also provided by estimates of O3-associated mortality 

and morbidity based on application of concentration-response relationships from epidemiologic 

studies to air quality adjusted to just meet the current standard. While these estimates are 

associated with uncertainties that complicate their interpretation, they suggest that O3-associated 

mortality and morbidity would be expected to decrease nationwide following reductions in 

precursor emissions to meet the current O3 standard. 

 

2

 HREA analyses for exposures of concern and for risk of moderate or large lung function decrements covered 15 

urban case study areas. HREA analyses of mortality and morbidity endpoints from epidemiologic studies covered 

12 urban case study areas. Exposures and risks were evaluated for the years 2006 through 2010. 
 ES-3

While meeting the current O3 standard is estimated to result in important public health 

improvements compared to recent air quality, staff further concludes that the O3-attributable 

health effects estimated to be allowed by air quality that meets the current primary standard can 

reasonably be judged important from a public health perspective. This conclusion is based on 

consideration of: (1) the scientific evidence discussed in the ISA, including controlled human 

exposure studies reporting abnormal or adverse respiratory effects following exposures to O3

concentrations below the level of the current standard and epidemiologic studies indicating 

associations with morbidity and mortality for air quality that would likely meet the current 

standard; (2) HREA estimates of O3 exposures of concern, O3-induced lung function risks, and 

O3-associated morbidity and mortality risks; (3) advice received from CASAC based on their 

review of draft versions of the ISA, HREA, and PA, and advice received in previous reviews; 

and (4) staff consideration of public comments. Staff reaches the overall conclusion that the 

available health evidence and exposure/risk information call into question the adequacy of the 

public health protection provided by the current standard. 

Given this conclusion regarding the adequacy of the current standard, staff also reaches 

conclusions for the Administrator’s consideration regarding the elements of potential alternative 

primary O3 standards that could be supported by the available evidence and exposure/risk 

information. Any such potential alternative standards should protect public health against effects 

associated with exposures to O3, alone or in combination with related photochemical oxidants, 

taking into account the available scientific evidence and exposure/risk information. In reaching 

conclusions about the range of potential alternative standards appropriate for consideration, staff 

is mindful that the Act requires primary standards that, in the judgment of the Administrator, are 

requisite to protect public health with an adequate margin of safety. In setting a primary standard 

that is “requisite” to protect public health, the EPA’s task is to establish standards that are neither 

more nor less stringent than necessary. The requirement that primary standards provide an 

“adequate margin of safety” is intended to address uncertainties associated with inconclusive 

scientific and technical information. Thus, the Act does not require that primary NAAQS be set 

at zero-risk levels, but rather at levels that reduce risk sufficiently to protect public health with an 

adequate margin of safety. 

The degree of public health protection provided by any NAAQS results from the 

collective impact of the elements of the standard, including the indicator, averaging time, form, 

and level. Staff’s conclusions on each of these elements are summarized below. 

(1) Indicator: It is appropriate to continue to use O3 as the indicator for a standard that is 

intended to address effects associated with exposure to O3, alone or in combination with 

related photochemical oxidants. Based on the available information, staff concludes that 
 ES-4

there is no basis for considering any alternative indicator at this time. Meeting an O3

standard can be expected to provide some degree of protection against potential health 

effects that may be independently associated with other photochemical oxidants, even 

though such effects are not discernible from currently available studies indexed by O3

alone. Staff notes that control of ambient O3 concentrations is generally understood to 

provide the best means of controlling photochemical oxidants, and thus of protecting 

against effects that may be associated with individual species and/or the broader mix of 

photochemical oxidants. CASAC concurred with these conclusions. 

(2) Averaging time: It is appropriate to consider retaining the current 8-hour averaging time 

for the primary O3 standard. 

(a) Staff concludes that an 8-hour averaging time remains appropriate for addressing 

health effects associated with short-term exposures to ambient O3. An 8-hour 

averaging time is similar to the exposure periods evaluated in controlled human 

exposure studies, including recent studies reporting respiratory effects following 

exposures to O3 concentrations below the level of the current standard. In 

addition, epidemiologic studies provide evidence for health effect associations 

with 8-hour O3 concentrations, as well as with 1-hour and 24-hour concentrations. 

Staff concludes that a standard with an 8-hour averaging time (combined with an 

appropriate standard form and level) would be expected to provide substantial 

protection against health effects attributable to 1- and 24-hour exposures. CASAC 

concurred, concluding that the current 8-hour averaging time is justified by the 

combined evidence from epidemiologic and clinical studies.

(b) Staff also concludes that a standard with an 8-hour averaging time can provide 

protection against respiratory effects associated with longer term O3 exposures. 

Air quality analyses indicate that just meeting an 8-hour standard with an 

appropriate level (i.e., 70 to 60 ppb, as discussed below) would be expected to 

maintain long-term O3 concentrations (i.e., seasonal average of 1-hour daily max) 

below those where a key study indicates the most confidence in the concentrationresponse relationship with respiratory mortality. In addition, risk analyses in the 

HREA estimate that just meeting such alternative 8-hour standards would be 

expected to decrease the incidence of respiratory mortality associated with longterm O3 concentrations. In considering other long-term O3 metrics evaluated in 

recent health studies, analyses in the HREA indicate that the large majority of the 

U.S. population lives in locations where reducing precursor emissions would be 
 ES-5

expected to decrease warm season averages of daily 8-hour ambient O3

concentrations, a long-term metric used in several recent studies reporting 

associations with respiratory morbidity. Taken together, these analyses suggest 

that a standard with an 8-hour averaging time, coupled with the current 4th-highest 

form and an appropriate level (discussed below), could provide appropriate 

protection against the long-term O3 concentrations reported to be associated with 

respiratory morbidity and mortality. CASAC concurred, concluding that the 8-

hour averaging time provides protection against the adverse impacts of long-term 

O3 exposures. 

(3) Form: For an 8-hour O3 standard with a revised level, as described below, it is 

appropriate to consider retaining the current form, defined as the 3-year average of the 

annual 4th-highest daily maximum concentration. Staff notes that this form was selected 

in 1997 and 2008 in recognition of the public health protection provided, when coupled 

with an appropriate averaging time and level, combined with the stability provided for 

implementation programs. The currently available evidence and exposure/risk 

information do not call into question these conclusions from previous reviews. CASAC 

concurred with this conclusion, agreeing that the current form, combined with the current 

8-hour averaging time, provides health protection while allowing for atypical 

meteorological conditions that can lead to abnormally high ambient O3 concentrations 

which, in turn, provides programmatic stability. 

(4) Level: The available scientific evidence and exposure/risk information provide strong 

support for considering a primary O3 standard with a revised level in order to increase 

public health protection, including for at-risk populations and lifestages. Staff concludes 

that it is appropriate in this review to consider a revised primary O3 standard level within 

the range of 70 ppb to 60 ppb. A standard set within this range would result in important 

improvements in public protection, compared to the current standard, and could 

reasonably be judged to provide an appropriate degree of public health protection, 

including for at-risk populations and lifestages. In its advice to the Administrator, 

CASAC also concluded that the scientific evidence and exposure/risk information 

support consideration of standard levels from 70 to 60 ppb. Within this range, CASAC 

concluded that a level of 70 ppb would provide little margin of safety and, therefore, 

provided the policy advice that the level of the O3 standard should be set below 70 ppb. 
 ES-6

The Administrator’s consideration of specific standard levels will reflect her judgments 

as to the appropriate weight to be given to various aspects of the scientific evidence and 

exposure/risk information, including the appropriate weight to be given to important 

uncertainties. To inform these judgments, staff considers what the evidence and 

information indicate with regard to the degree of public health protection that could be 

achieved with levels from the upper (70 ppb), middle (65 ppb), and lower (60 ppb) parts 

of the range. 

A level of 70 ppb is below the O3 exposure concentration that has been reported to elicit a 

broad range of respiratory effects that includes airway hyperresponsiveness and 

decreased lung host defense, in addition to lung function decrements, airway 

inflammation, and respiratory symptoms (i.e., 80 ppb). A level of 70 ppb is also just 

below the lowest exposure concentration at which the combined occurrence of respiratory 

symptoms and lung function decrements have been reported (i.e., 72 ppb), a combination 

judged adverse by the ATS (section 3.1.3). A level of 70 ppb is above the lowest 

exposure concentration demonstrated to result in lung function decrements and 

pulmonary inflammation (i.e., 60 ppb). Compared to the current standard, a revised O3

standard with a level of 70 ppb would be expected to (1) reduce the occurrence of 

exposures of concern to O3 concentrations that result in respiratory effects in healthy 

adults (at or above 60 and 70 ppb) by about 45 to 95%, almost eliminating the 

occurrence of multiple exposures at or above 70 ppb; (2) reduce the occurrence of 

moderate-to-large O3-induced lung function decrements (FEV1 decrements > 10, 15, 

20%) by about 15 to 35%, most effectively limiting the occurrence of multiple 

decrements and decrements > 15, 20%; (3) more effectively maintain short- and longterm O3 concentrations below those present in the epidemiologic studies that reported 

significant O3 health effect associations in locations likely to have met the current 

standard;3

 and (4) reduce the risk of O3-associated mortality and morbidity, particularly 

the risk associated with the upper portions of the distributions of ambient O3

concentrations. 

A level of 65 ppb is well below the O3 exposure concentration that has been reported to 

elicit the wide range of potentially adverse respiratory effects noted above, and is below 

the lowest exposure concentration at which the combined occurrence of respiratory 

 

3

 Though epidemiologic studies also provide evidence for O3 health effect associations in locations likely to have 

met a standard with a level of 70 ppb, as discussed below for lower standard levels. 
 ES-7

symptoms and lung function decrements has been reported. As noted above for 70 ppb, a 

level of 65 ppb is above the lowest exposure concentration demonstrated to result in lung 

function decrements and pulmonary inflammation. Compared to a standard with a level 

of 70 ppb, a revised standard with a level of 65 ppb would be expected to (1) further 

reduce the occurrence of exposures of concern (by about 80 to 100% compared to the 

current standard), decreasing exposures at or above 60 ppb and almost eliminating 

exposures at or above 70 and 80 ppb; (2) further reduce the occurrence of FEV1

decrements > 10, 15, and 20% (by about 30 to 65%, compared to the current standard); 

(3) more effectively maintain short- and long-term O3 concentrations below those present 

in the epidemiologic studies that reported significant O3 health effect associations in 

locations likely to have met the current standard;4

 and (4) further reduce the risk of O3-

associated mortality and morbidity, particularly the risk associated with the upper portion 

of the distribution of ambient O3 concentrations. 

A level of 60 ppb is well below the O3 exposure concentration shown to result in the 

combined occurrence of respiratory symptoms and lung function decrements, and 

corresponds to the lowest exposure concentration demonstrated to result in lung function 

decrements and pulmonary inflammation. Compared to a standard with a level of 70 or 

65 ppb, a revised standard with a level of 60 ppb would be expected to (1) further reduce 

the occurrence of exposures of concern (by about 95 to 100% compared to the current 

standard), almost eliminating exposures at or above 60 ppb; (2) further reduce the 

occurrence of FEV1 decrements > 10, 15, and 20%, (by about 45 to 85% compared to the 

current standard); (3) more effectively maintain short- and long-term O3 concentrations 

below those present in the epidemiologic studies that reported significant O3 health effect 

associations in locations likely to have met the current standard;5 and (4) further reduce 

the risk of O3-associated mortality and morbidity, particularly the risk associated with the 

upper portion of the distribution of ambient O3 concentrations. 

Welfare Effects and Review of the Secondary Standard 

The longstanding and comprehensive evidence base, stronger than in the last review, 

documents the vegetation and ecosystem-related effects of O3 in ambient air. In particular, recent 

controlled studies at the molecular, biochemical and cellular scales have increased the 

 

4

 Though epidemiologic studies also provide evidence for O3 health effect associations in locations likely to have 

met a standard with a level of 65 ppb. 

5

 Epidemiologic studies have not evaluated O3 health effect associations based primarily on air quality in locations 

likely to have met a standard with a level of 60 ppb. 
 ES-8

mechanistic understanding of the basic biology of how plants are affected by oxidative stress. 

These studies have focused on a variety of plant responses to O3 including: 1) reduced carbon 

dioxide uptake due to stomatal closure; 2) the upregulation of genes associated with plant 

defense, signaling, hormone synthesis and secondary metabolism; 3) the down regulation of 

genes related to photosynthesis and general metabolism; 4) the loss of carbon assimilation 

capacity due to declines in the quantity and activity of key proteins and enzymes; and 5) the 

negative impacts on the efficiency of the photosynthetic light reactions. In addition, these effects 

at the plant scale can be linked to an array of effects at larger scales, as shown in recent field 

studies, together with previously available evidence. Specifically, plant-scale effects, such as 

altered rates of leaf gas exchange, growth, and reproduction at the individual plant level, can 

result in larger scale effects in ecosystems, such as alterations in productivity, carbon storage, 

water cycling, nutrient cycling, and community composition. The available information also 

demonstrates a relationship between changes in tropospheric O3 concentrations and radiative 

forcing, and between changes in tropospheric O3 concentrations and effects on climate. 

The long-standing body of available evidence also provides a wealth of information on 

aspects of O3 exposure that are important in influencing plant response. These include support 

for the conclusions that: O3 effects in plants are cumulative; higher O3 concentrations appear to 

be more important than lower concentrations in eliciting a response; plant sensitivity to O3 varies 

with time of day and plant development stage; and quantifying exposure with indices that 

cumulate hourly O3 concentrations and preferentially weight the higher concentrations improves 

the explanatory power of exposure/response models for growth and yield, over using indices 

based on mean and peak exposure values. 

As an initial matter in this PA, staff concludes that reducing ambient O3 concentrations to 

meet the current standard of 75 ppb will provide important improvements in public welfare 

protection. This initial conclusion is based on (1) the strong body of scientific evidence 

indicating a wide range of effects to sensitive vegetation, including tree biomass loss, crop yield 

loss, and visible foliar injury, and associated ecosystems and services attributable to cumulative 

exposures to O3 concentrations found in the ambient air and (2) estimates indicating decreased 

cumulative O3 exposures and welfare risks upon meeting the current standard, compared to 

recent air quality. Strong support for this conclusion is provided by the available welfare 

evidence; by WREA estimates of cumulative exposures to O3 concentrations shown to result in 

decreased biomass loss, crop yield loss, and visible foliar injury incidence under just meeting the 

current secondary standard; and by WREA estimates of improvements in carbon storage and air 

pollution removal in urban and commercial forests. Support for this conclusion is also provided 

by WREA estimates of increased protection for Class I areas from O3-associated visible foliar 

injury and tree biomass loss. 
 ES-9

Staff further concludes that the O3-attributable welfare effects estimated to be allowed by 

air quality that meets the current secondary standard call into question the adequacy of the public 

welfare protection provided by the current standard. In addition, staff also concludes that the 

public welfare protection is most appropriately judged through the use of a more biologically 

relevant form, such as the cumulative, seasonal W126-metric. These conclusions are based on 

consideration of: (1) the scientific evidence, including controlled exposure studies reporting 

effects on plant growth, productivity and carbon storage, crop yield loss, and visible foliar injury 

following exposures to O3 concentrations below the level of the current standard and field based 

studies that support these conclusions for air quality that would likely meet the current standard; 

(2) the longstanding and extensive evidence demonstrating that the risk to vegetation comes from 

cumulative seasonal exposures; (3) evidence suggesting that in Class I areas meeting the current 

standard, cumulative seasonal O3 exposures occur that are associated with estimates of tree 

growth impacts of a magnitude that are reasonably considered important to public welfare; (4) 

WREA estimates of reductions in biomass loss, crop yield loss, and visible foliar injury 

incidence, and improvements in carbon storage and air pollution removal in urban and 

commercial forests when meeting alternative W126 levels; (5) advice received from CASAC 

based on their review of draft versions of the ISA, WREA, and PA, and advice received in 

previous reviews; and (6) public comments. Staff reaches the overall conclusion that the 

available vegetation and ecosystem effects evidence and exposure/risk information, including for 

associated ecosystem services important from a public welfare perspective, call into question the 

adequacy of the public welfare protection provided by the current standard. Based on the 

evaluation presented in this PA, staff concludes that consideration should be given to revising the 

standard to provide increased public welfare protection. CASAC agreed with this conclusion. 

Given this conclusion regarding the adequacy of the current standard, staff also reaches 

conclusions for the Administrator’s consideration regarding the elements of potential alternative 

secondary O3 standards that could be supported by the available evidence and exposure/risk 

information. Any such potential alternative standards should protect public welfare against 

known or anticipated adverse environmental effects associated with exposures to O3, alone or in 

combination with related photochemical oxidants, taking into account the available scientific 

evidence and exposure/risk information. In reaching conclusions about the range of potential 

alternative standards appropriate for consideration, staff is mindful that the Act requires 

secondary standards that are at “a level of air quality the attainment and maintenance of which” 

in the “judgment of the Administrator”, are “requisite to protect public welfare from any known 

or anticipated adverse effects”. In setting a secondary standard that is “requisite” to protect 

public welfare, the EPA’s task is to establish standards that are neither more nor less stringent 
 ES-10

than necessary. Thus, the Act does not require that NAAQS be set at zero-risk levels, but rather 

at levels that reduce risk sufficiently to protect public welfare from adverse effects. 

The degree of public welfare protection provided by any NAAQS results from the 

collective impact of the elements of the standard, including the indicator, averaging time, form, 

and level. Staff’s conclusions on each of these elements are summarized below. 

(1) Indicator: Staff concludes that it is appropriate to continue to use O3 as the indicator 

for a standard that is intended to address welfare effects associated with exposure to 

O3, alone or in combination with related photochemical oxidants. Based on the 

available information, staff concludes that there is no basis for considering an 

alternative indicator at this time. CASAC concurred with these conclusions. 

(2) Averaging time and form: Staff concludes that it is appropriate to consider a revised 

secondary standard in terms of the cumulative, seasonal, concentration-weighted 

form, called the W126 index. This is supported by strong scientific evidence that 

cumulative O3 exposures drive plant response and can cause reduced tree growth, 

productivity, and carbon storage; crop yield loss; visible foliar injury; and other 

welfare effects. With regard to the appropriate form and averaging times, staff 

reaches the following additional conclusions: 

(a) It is appropriate to consider the consecutive 3-month period within the O3

season with the maximum index value as the seasonal period over which 

to cumulate hourly O3 exposures. Staff notes that the maximum 3-month 

period generally coincides with maximum biological activity for most 

vegetation, making the 3-month duration a suitable surrogate for longer 

growing seasons. 

(b) It is appropriate to cumulate daily exposures for the 12-hour period from 

8:00 am to 8:00 pm, generally representing the daylight period during the 

3-month period identified above. 

To the extent the Administrator finds it useful to consider the extent of public welfare 

protection that might be afforded by a revised primary standard, staff concludes that 

public welfare protection is appropriately judged through the use of the cumulative, 

seasonal W126 index form, as described above. CASAC agreed that it was 
 ES-11

appropriate to establish a revised form of the secondary standard and that the W126 

index was a more biologically relevant form than the current form of the standard. 

 With regard to the number of years over which it is appropriate to average, staff 

notes the that there is limited information to discern between the level of protection 

provided by an annual form or a 3-year average form of a W126 standard for crop 

yield loss or foliar injury, and that a multiple year form could be considered to 

provide a more consistent target level of protection for this endpoint. Such a form 

might also be appropriate for a standard intended to achieve the desired level of 

protection from longer-term effects, including those associated with potential 

compounding of biomass loss over multiple years. Further, such a form might be 

concluded to contribute to greater stability in air quality management programs, and 

thus, greater effectiveness in achieving the desired level of public welfare protection, 

than that that might result from a single year form. Therefore, to the extent that the 

greater emphasis is placed on protecting against effects associated with multi-year 

exposures and maintaining more year-to-year stability of public welfare protection, 

staff concludes that it is appropriate to consider a secondary standard form that 

averages the seasonal W126 index values across three consecutive years. CASAC 

recommended that if a 3-year averaging period is selected, the level should be set 

lower than if a 1-year averaging period is selected in order to provide greater 

protection for annual crops and against cumulative effects on perennial species. 

(3) Level: With regard to level for a revised secondary standard, staff concludes that it is 

appropriate to give consideration to a range of levels from 17 to 7 ppm-hrs, expressed 

in terms of the W126 index. In so doing, we primarily consider the evidence- and 

exposure/risk-based information for cumulative seasonal O3 exposures represented by 

W126 index values (including those represented by the WREA average W126 

scenarios) associated with biomass loss in studied tree species, both in and outside 

areas that have been afforded special protections. We note CASAC’s advice that a 6% 

median RBL is unacceptably high, and that the 2% median RBL is an important 

benchmark to consider. We further note that for the lower level of 7 ppm-hrs the 

median tree species biomass loss is at or below 2% and that for the upper level of 17 

ppm-hrs the median tree biomass loss is below 6%.6

 We also note that a level of 17 

ppm-hrs reduces the percent of total area having weighted RBL greater than 2% to 

 

6

 We note that a W126 index value of 19 ppm-hrs is estimated to result in a median RBL value of 6%. 
 ES-12

0.2%, and reduces the number of Class I areas with weighted RBL greater than 2% to 

2 of the 145 assessed nationally protected Class I areas. 

We also note that tree biomass loss can be an indicator of more significant ecosystemwide effects which might reasonably be concluded to be significant to public welfare. 

For example, when it occurs over multiple years at a sufficient magnitude, biomass 

loss is linked to an array of effects on other ecosystem-level processes such as 

nutrient and water cycles, changes in above and below ground communities, and 

carbon storage and air pollution removal. These effects have the potential to be 

adverse to the public welfare. 

In addition, a range of levels from 17 to 7 ppm-hrs would protect at least half of the 

crop species from a yield loss of greater than 5%. A W126 level of 10 ppm-hrs or less 

would also reduce prevalence of visible foliar injury and promote appreciable gains in 

carbon sequestration and pollutant removal. 

CASAC recommended a range of W126 values of 15 ppm-hrs to 7 ppm-hrs and did 

not recommend levels above 15 ppm-hrs. CASAC noted that a level of 15 ppm-hrs is 

requisite to protect median crop yield loss to no more than 5% and that a level below 

10 ppm-hrs is required to reduce foliar injury prevalence. CASAC also noted that a 

W126 level of 7 ppm-hrs limits median relative biomass loss for trees to no greater 

than 2% and offers additional protection against crop yield loss and foliar injury. 

The Administrator’s consideration of a particular level within the range of 17 to 7 ppmhrs would reflect judgments as to the appropriate weight to be given to various aspects of the 

scientific evidence and exposure/risk information, with appropriate weight given to important 

uncertainties and with particular consideration of the support provided by this evidence and 

information regarding the protection of public welfare. To the extent the Administrator finds it 

useful to consider the extent of public welfare protection that might be afforded by a revised 

primary standard, staff concludes that public welfare protection is appropriately judged through 

the use of the cumulative seasonal W126-based metric. 
1-1 

1 INTRODUCTION 

1.1 PURPOSE 

The U.S. Environmental Protection Agency (EPA) is presently conducting a review of 

the primary (health-based) and secondary (welfare-based) national ambient air quality standards 

(NAAQS) for ozone (O3). The overall plan for this review was presented in the Integrated 

Review Plan for the O3 National Ambient Air Quality Standards (IRP, U.S. EPA, 2011a). The 

IRP also identified key policy-relevant issues to be addressed in this review and discussed the 

key documents that generally inform NAAQS reviews, including an Integrated Science 

Assessment (ISA), Risk and Exposure Assessments (REAs), and a Policy Assessment (PA). The 

PA is prepared by the staff in EPA’s Office of Air Quality Planning and Standards (OAQPS). It 

presents a staff evaluation of the policy implications of the key scientific and technical 

information in the ISA and REAs for EPA’s consideration.1 The PA provides a transparent 

evaluation, and staff conclusions, regarding policy considerations related to reaching judgments 

about the adequacy of the current standards, and if revision is considered, what revisions may be 

appropriate to consider. 

The PA is intended to help “bridge the gap” between the Agency’s scientific assessments 

presented in the ISA and REAs, and the judgments required of the EPA Administrator in 

determining whether it is appropriate to retain or revise the NAAQS.2 In evaluating the 

adequacy of the current standard and whether it is appropriate to consider potential alternative 

standards, the PA focuses on information that is most pertinent to evaluating the basic elements 

of the NAAQS: indicator,3

 averaging time, form,4 and level. These elements, which together 

serve to define each standard, must be considered collectively in evaluating the health and 

welfare protection afforded by the O3 standards. The PA integrates and interprets the information 

from the ISA and REAs to frame policy options for consideration by the Administrator. In so 

doing, the PA recognizes that the selection of a specific approach to reaching final decisions on 

the primary and secondary O3 standards will reflect the judgments of the Administrator. 

 

1

The terms “staff” and “we” through this document refer to personnel in the EPA’s Office of Air Quality Planning 

and Standards (OAQPS). 

2

American Farm Bureau Federation v. EPA, 559 F. 3d 512, 521 (D.C. Cir. 2009); Natural Resources Defense 

Council v. EPA, 902 F. 2d 962, 967-68, 970 (D.C. Cir. 1990). 

3

The “indicator” of a standard defines the chemical species or mixture that is to be measured in determining whether 

an area attains the standard. The indicator for photochemical oxidants is ozone. 

4

The “form” of a standard defines the air quality statistic that is to be compared to the level of the standard in 

determining whether an area attains the standard. For example, the form of the current 8-hour O3 NAAQS is the 3-

year average of the annual fourth-highest daily maximum 8-hour average. 
1-2 

The development of the PA is also intended to facilitate advice to the Agency and 

recommendations to the Administrator from an independent scientific review committee, the 

Clean Air Scientific Advisory Committee (CASAC), as provided for in the Clean Air Act. As 

discussed below in section 1.2.1, the CASAC is to advise not only on the Agency’s assessment 

of the relevant scientific information, but also on the adequacy of the existing standards, and to 

make recommendations as to any revisions of the standards that may be appropriate. The EPA 

facilitates CASAC advice and recommendations, as well as public input and comment, by 

requesting CASAC review and public comment on one or more drafts of the PA. 

In this PA for the review of the O3 NAAQS, we5 consider the scientific and technical 

information available in this review as assessed in the Integrated Science Assessment for O3 and 

Related Photochemical Oxidants (ISA, U.S. EPA, 2013), prepared by EPA’s National Center for 

Environmental Assessment (NCEA), and the quantitative human exposure and health risk 

assessment and welfare risk assessment documents (HREA, U.S. EPA, 2014a; WREA, U.S. 

EPA, 2014b). The evaluation and staff conclusions presented in this PA have been informed by 

comments and advice received from CASAC in their reviews of draft versions of the PA, and in 

their reviews of the other draft Agency documents prepared for this NAAQS review. 

Beyond informing the EPA Administrator and facilitating the advice and 

recommendations of CASAC and the public, the PA is also intended to be a useful reference to 

all parties interested in the NAAQS review. In these roles, it is intended to serve as a single 

source of the most policy-relevant information that informs the Agency’s review of the NAAQS, 

and it is written to be understandable to a broad audience. 

The remainder of chapter 1 summarizes information on the NAAQS legislative 

requirements and on the history of the O3 NAAQS (section 1.2), and summarizes our general 

approaches to reviewing the current O3 NAAQS (section 1.3). Chapter 2 of this PA provides an 

overview of the O3 ambient monitoring network and O3 air quality, including estimates of O3

concentrations attributable to background sources. The remaining chapters are organized into 

two main parts. Chapters 3 and 4 focus on the review of the primary O3 NAAQS while chapters 

5 and 6 focus on the review of the secondary O3 NAAQS. Staff’s considerations and conclusions 

related to the current primary and secondary standards are discussed in chapters 3 and 5, 

respectively. Staff’s considerations and conclusions related to potential alternative primary and 

secondary standards are discussed in chapters 4 and 6, respectively. Key uncertainties in the 

review and areas for future research and data collection are additionally identified in chapters 4 

and 6 for the two types of standards. 

 

5

As noted above, the term “we” through this document refer to personnel in the EPA’s Office of Air Quality 

Planning and Standards (OAQPS). 
1-3 

1.2 BACKGROUND 

1.2.1 Legislative Requirements 

Two sections of the Clean Air Act (CAA) govern the establishment and revision of the 

NAAQS. Section 108 (42 U.S.C. section 7408) directs the Administrator to identify and list 

certain air pollutants and then to issue air quality criteria for those pollutants. The Administrator 

is to list those air pollutants that in her “judgment, cause or contribute to air pollution which may 

reasonably be anticipated to endanger public health or welfare;” “the presence of which in the 

ambient air results from numerous or diverse mobile or stationary sources;” and “for which . . . 

[the Administrator] plans to issue air quality criteria....” Air quality criteria are intended to 

“accurately reflect the latest scientific knowledge useful in indicating the kind and extent of all 

identifiable effects on public health or welfare which may be expected from the presence of [a] 

pollutant in the ambient air . . .” 42 U.S.C. § 7408(b). Section 109 (42 U.S.C. 7409) directs the 

Administrator to propose and promulgate “primary” and “secondary” NAAQS for pollutants for 

which air quality criteria are issued. Section 109(b)(1) defines a primary standard as one “the 

attainment and maintenance of which in the judgment of the Administrator, based on such 

criteria and allowing an adequate margin of safety, are requisite to protect the public health.” 6 

A secondary standard, as defined in section 109(b)(2), must “specify a level of air quality the 

attainment and maintenance of which, in the judgment of the Administrator, based on such 

criteria, is requisite to protect the public welfare from any known or anticipated adverse effects 

associated with the presence of [the] pollutant in the ambient air.” 7 

The requirement that primary standards provide an adequate margin of safety was 

intended to address uncertainties associated with inconclusive scientific and technical 

information available at the time of standard setting. It was also intended to provide a reasonable 

degree of protection against hazards that research has not yet identified. See State of Mississippi 

v. EPA, 744 F. 3d 1334, 1353 (D.C. Cir. 2012) (“By requiring an ‘adequate margin of safety’, 

Congress was directing EPA to build a buffer to protect against uncertain and unknown dangers 

to human health”). See also Lead Industries Association v. EPA, 647 F.2d 1130, 1154 (D.C. Cir 

1980); American Petroleum Institute v. Costle, 665 F.2d 1176, 1186 (D.C. Cir. 1981); American 

Farm Bureau Federation v. EPA, 559 F. 3d 512, 533 (D.C. Cir. 2009); Association of Battery 

 

6

 The legislative history of section 109 indicates that a primary standard is to be set at “the maximum permissible 

ambient air level . . . which will protect the health of any [sensitive] group of the population,” and that for this 

purpose “reference should be made to a representative sample of persons comprising the sensitive group rather than 

to a single person in such a group” S. Rep. No. 91-1196, 91st Cong., 2d Sess. 10 (1970). 

7

 Welfare effects as defined in section 302(h) (42 U.S.C. § 7602(h)) include, but are not limited to, “effects on soils, 

water, crops, vegetation, man-made materials, animals, wildlife, weather, visibility and climate, damage to and 

deterioration of property, and hazards to transportation, as well as effects on economic values and on personal 

comfort and well-being.” 
1-4 

Recyclers v. EPA, 604 F. 3d 613, 617-18 (D.C. Cir. 2010). Both kinds of uncertainties are 

components of the risk associated with pollution at levels below those at which human health 

effects can be said to occur with reasonable scientific certainty. Thus, in selecting primary 

standards that provide an adequate margin of safety, the Administrator is seeking not only to 

prevent pollution levels that have been demonstrated to be harmful but also to prevent lower 

pollutant levels that may pose an unacceptable risk of harm, even if the risk is not precisely 

identified as to nature or degree. The CAA does not require the Administrator to establish a 

primary NAAQS at a zero-risk level or at background concentration levels, see Lead Industries 

v. EPA, 647 F.2d at 1156 n.51; State of Mississippi v. EPA, 744 F. 3d at 1343, 1351, but rather 

at a level that reduces risk sufficiently so as to protect public health with an adequate margin of 

safety. 

In addressing the requirement for an adequate margin of safety, the EPA considers such 

factors as the nature and severity of the health effects, the size of sensitive population(s)8 at risk, 

and the kind and degree of the uncertainties that must be addressed. The selection of any 

particular approach for providing an adequate margin of safety is a policy choice left specifically 

to the Administrator’s judgment. See Lead Industries Association v. EPA, 647 F.2d at 1161-62; 

State of Mississippi, 744 F. 3d at 1353. 

In setting primary and secondary standards that are “requisite” to protect public health 

and welfare, respectively, as provided in section 109(b), EPA’s task is to establish standards that 

are neither more nor less stringent than necessary for these purposes. In so doing, the EPA may 

not consider the costs of implementing the standards. See generally, Whitman v. American 

Trucking Associations, 531 U.S. 457, 465-472, 475-76 (2001). Likewise, “[a]ttainability and 

technological feasibility are not relevant considerations in the promulgation of national ambient 

air quality standards.” American Petroleum Institute v. Costle, 665 F. 2d at 1185. 

Section 109(d)(1) requires that “not later than December 31, 1980, and at 5-year intervals 

thereafter, the Administrator shall complete a thorough review of the criteria published under 

section 108 and the national ambient air quality standards . . . and shall make such revisions in 

such criteria and standards and promulgate such new standards as may be appropriate . . . .” 

Section 109(d)(2) requires that an independent scientific review committee “shall complete a 

review of the criteria . . . and the national primary and secondary ambient air quality standards . . 

. and shall recommend to the Administrator any new . . . standards and revisions of existing 

 

8

 As used here and similarly throughout this document, the term population refers to persons having a quality or 

characteristic in common, including a specific pre-existing illness or a specific age or life stage. 
1-5 

criteria and standards as may be appropriate . . . .” Since the early 1980's, the Clean Air 

Scientific Advisory Committee (CASAC) has performed this independent review function. 9 

1.2.2 History of O3 NAAQS Reviews 

Table 1-1 summarizes the O3 NAAQS that the EPA has promulgated to date. In each 

review, the EPA set the secondary standard at a level identical to the primary standard. These 

reviews are briefly described below. 

Table 1-1. Summary of primary and secondary O3 NAAQS promulgated during the 

period from 1971 to 2008. 

Final Rule Indicator Averaging Time Level (ppm) Form 

1971 (36 FR 8186) Total photochemical 

oxidants 1 hour 0.08 Not to be exceeded more than 

one hour per year 

1979 (44 FR 8202) 03 1 hour 0.12 

Attainment is defined when the 

expected number of days per 

calendar year, with maximum 

hourly average concentration 

greater than 0.12 ppm, is 

equal to or less than 1 

1993 (58 FR 13008) The EPA decided that revisions to the standards were not warranted at the time. 

1997 (62 FR 38856) 03 8 hours 0.08 

Annual fourth-highest daily 

maximum 8-hour 

concentration, averaged over 3 

years 

2008 (73 FR 16483) 03 8 hours 0.075 

Form of the standards 

remained unchanged relative 

to the 1997 standard 

The EPA first established primary and secondary NAAQS for photochemical oxidants in 

1971 (36 FR 8186, April 30, 1971). The EPA set both primary and secondary standards at a level 

of 0.08 parts per million (ppm), 1-hr average, total photochemical oxidants, not to be exceeded 

more than one hour per year. The EPA based the standards on scientific information contained in 

the 1970 Air Quality Criteria for Photochemical Oxidants (U.S. DHEW, 1970). We initiated the 

first periodic review of the NAAQS for photochemical oxidants in 1977. Based on the 1978 Air 

 

9 Lists of CASAC members and of members of the CASAC Ozone Review Panel are available at: 

http://yosemite.epa.gov/sab/sabpeople.nsf/WebCommitteesSubCommittees/Ozone%20Review%20Panel. 
1-6 

Quality Criteria for Ozone and Other Photochemical Oxidants (U.S. EPA, 1978), the EPA 

published proposed revisions to the original NAAQS in 1978 (43 FR 16962) and final revisions 

in 1979 (44 FR 8202). At that time, the EPA revised the level of the primary and secondary 

standards from 0.08 to 0.12 ppm and changed the indicator from photochemical oxidants to O3, 

and the form of the standards from a deterministic to a statistical form. This statistical form 

defined attainment of the standards as occurring when the expected number of days per calendar 

year with maximum hourly average concentration greater than 0.12 ppm equaled one or less. 

Following the final decision in the 1979 review, the City of Houston challenged the 

Administrator’s decision arguing that the standard was arbitrary and capricious because natural 

O3 concentrations and other physical phenomena in the Houston area made the standard 

unattainable in that area. The U.S. Court of Appeals for the District of Columbia Circuit (D.C. 

Circuit) rejected this argument, holding (as noted above) that attainability and technological 

feasibility are not relevant considerations in the promulgation of the NAAQS. The court also 

noted that the EPA need not tailor the NAAQS to fit each region or locale, pointing out that 

Congress was aware of the difficulty in meeting standards in some locations and had addressed 

this difficulty through various compliance related provisions in the Act. See API v. Costle, 665 

F.2d 1176, 1184-6 (D.C. Cir. 1981). 

In 1982, we announced plans to revise the 1978 Air Quality Criteria document (47 FR 

11561), and in 1983, we initiated the second periodic review of the O3 NAAQS (48 FR 38009). 

We subsequently published the 1986 Air Quality Criteria for Ozone and Other Photochemical 

Oxidants (U.S. EPA, 1986) and the 1989 Staff Paper (U.S. EPA, 1989). Following publication of 

the 1986 Air Quality Criteria Document (AQCD), a number of scientific abstracts and articles 

were published that appeared to be of sufficient importance concerning potential health and 

welfare effects of O3 to warrant preparation of a Supplement. On August 10, 1992, under the 

terms of a court order, the EPA published a proposed decision to retain the existing primary and 

secondary standards (57 FR 35542). The notice explained that the proposed decision would 

complete EPA’s review of information on health and welfare effects of O3 assembled over a 7-

year period and contained in the 1986 AQCD and its 1992 Supplement. The proposal also 

announced EPA’s intention to proceed as rapidly as possible with the next review of the air 

quality criteria and standards for O3 in light of emerging evidence of health effects related to 6- 

to 8-hour O3 exposures. On March 9, 1993, the EPA concluded the review by affirming its 

proposed decision to retain the existing primary and secondary standards. (58 FR 13008). 

In August 1992, we announced plans to initiate the third periodic review of the air quality 

criteria and O3 NAAQS (57 FR 35542). In December 1996, the EPA proposed to replace the then 

existing 1-hour primary and secondary standards with 8-hour average O3 standards set at a level 

of 0.08 ppm (equivalent to 0.084 ppm using standard rounding conventions) (61 FR 65716). The 
1-7 

EPA also proposed to establish a new distinct secondary standard using a biologically-based 

cumulative, seasonal form. The EPA completed this review on July 18, 1997 (62 FR 38856) by 

setting the primary standard at a level of 0.08 ppm, based on the annual fourth-highest daily 

maximum 8-hr average concentration, averaged over three years, and setting the secondary 

standard identical to the revised primary standard. In reaching this decision, the EPA identified 

several reasons supporting its decision to reject a potential alternate standard set at 0.07 ppm. 

Most importantly, the EPA pointed out the scientific uncertainty at lower concentrations and 

placed significant weight on the fact that no CASAC panel member supported a standard level 

set lower than 0.08 ppm (62 FR 38868). In addition to noting the uncertainties in the health 

evidence for exposure concentrations below 0.08 ppm and the advice of CASAC, the EPA noted 

that a standard set at a level of 0.07 ppm would be closer to peak background concentrations that 

infrequently occur in some areas due to nonanthropogenic sources of O3 precursors (62 FR 

38856, 38868; July 18, 1997). 

On May 14, 1999, in response to challenges by industry and others to EPA’s 1997 

decision, the U.S. Court of Appeals for the District of Columbia Circuit remanded the O3

NAAQS to the EPA, finding that section 109 of the Act, as interpreted by the EPA, effected an 

unconstitutional delegation of legislative authority. American Trucking Assoc. vs. EPA, 175 

F.3d 1027, 1034-1040(D.C. Cir. 1999) (“ATA I”). In addition, the court directed that, in 

responding to the remand, the EPA should consider the potential beneficial health effects of O3

pollution in shielding the public from the effects of solar ultraviolet (UV) radiation, as well as 

adverse health effects. Id. at 1051-53. In 1999, the EPA petitioned for rehearing en banc on 

several issues related to that decision. The court granted the request for rehearing in part and 

denied it in part, but declined to review its ruling with regard to the potential beneficial effects of 

O3 pollution. 195 F3d 4, 10 (D.C Cir., 1999) (“ATA II”). On January 27, 2000, the EPA 

petitioned the U.S. Supreme Court for certiorari on the constitutional issue (and two other 

issues), but did not request review of the 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. Whitman v. American Trucking Assoc., 531 U. S. 457, 

472-74 (2001) (holding that section 109 of the CAA does not delegate legislative power to the 

EPA in contravention of the Constitution). The Court 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 

decisions. On March 26, 2002, the D.C. Circuit issued its final decision on remand, finding the 

1997 O3 NAAQS to be “neither arbitrary nor capricious,” and so denying the remaining petitions 

for review. American Trucking Associations, Inc. v EPA, 283 F.3d 355, 379 (D.C Cir., 2002) 

(“ATA III”). 
1-8 

Specifically, in ATA III, the D.C. Circuit upheld EPA’s decision on the 1997 O3 standard 

as the product of reasoned decision-making. The Court made clear that the most important 

support for EPA’s decision was the health evidence and the concerns it raised about setting a 

standard level below 0.08 ppm. (“the record is replete with references to studies demonstrating 

the inadequacies of the old one-hour standard”, as well as extensive information supporting the 

change to an 8-hour averaging time). 283 F 3d at 378. The Court also pointed to the significant 

weight that the EPA properly placed on the advice it received from CASAC. Id. at 379. The 

court further noted that “although relative proximity to peak background ozone concentrations 

did not, in itself, necessitate a level of 0.08, EPA could consider that factor when choosing 

among the three alternative levels.” Id. 

Independently of the litigation, the EPA also responded to the Court’s remand to 

consider the potential beneficial health effects of O3 pollution in shielding the public from effects 

of solar (ultraviolet or UV-B) radiation. The EPA provisionally determined that the information 

linking changes in patterns of ground-level O3 concentrations to changes in relevant patterns of 

exposures to ultraviolet (UV-B) radiation of concern to public health was too uncertain, at that 

time, to warrant any relaxation in 1997 O3 NAAQS. The EPA also expressed the view that any 

plausible changes in UV-B radiation exposures from changes in patterns of ground-level O3

concentrations would likely be very small from a public health perspective. In view of these 

findings, the EPA proposed to leave the 1997 8-hour NAAQS unchanged (66 FR 57268, Nov. 

14, 2001). After considering public comment on the proposed decision, the EPA published its 

final response to this remand on January 6, 2003, re-affirming the 8-hour O3 NAAQS set in 1997 

(68 FR 614). 

The EPA initiated the fourth periodic review of the air quality criteria and O3 standards in 

September 2000 with a call for information (65 FR 57810). The schedule for completion of that 

review was ultimately governed by a consent decree resolving a lawsuit filed in March 2003 by 

plaintiffs representing national environmental and public health organizations, who maintained 

that EPA was in breach of a mandatory legal duty to complete review of the O3 NAAQS within a 

statutorily-mandated deadline. On July 11, 2007, the EPA proposed to revise the level of the 

primary standard within a range of 0.075 to 0.070 ppm. (72 FR 37818). Documents supporting 

this proposed decision included the Air Quality Criteria for Ozone and Other Photochemical 

Oxidants (U.S. EPA, 2006) and the Staff Paper (U.S EPA, 2007) and related technical support 

documents. The EPA also proposed two options for revising the secondary standard: (1) replace 

the current standard with a cumulative, seasonal standard, expressed as an index of the annual 

sum of weighted hourly concentrations cumulated over 12 daylight hours during the consecutive 

3-month period within the O3 season with the maximum index value, set at a level within the 

range of 7 to 21 ppm-hrs, and (2) set the secondary standard identical to the proposed primary 
1-9 

standard. The EPA completed the review with publication of a final decision on March 27, 2008 

(73 FR 16436). In that final rule, the EPA revised the NAAQS by lowering the level of the 8-

hour primary O3 standard from 0.08 ppm to 0.075 ppm, not otherwise revising the primary 

standard, and adopting a secondary standard identical to the revised primary standard. In May 

2008, state, public health, environmental, and industry petitioners filed suit challenging EPA’s 

final decision on the 2008 O3 standards. On September 16, 2009, the EPA announced its 

intention to reconsider the 2008 O3 standards, and initiated a rulemaking to do so. At EPA’s 

request, the Court held the consolidated cases in abeyance pending EPA’s reconsideration of the 

2008 decision. 

On January 19, 2010 (75 FR 2938), the EPA issued a notice of proposed rulemaking to 

reconsider the 2008 final decision. In that notice, the EPA proposed that further revisions of the 

primary and secondary standards were necessary to provide a requisite level of protection to 

public health and welfare. The EPA proposed to decrease the level of the 2008 8-hour primary 

standard from 0.075 ppm to a level within the range of 0.060 to 0.070 ppm, and to change the 

secondary standard to a new cumulative, seasonal standard expressed as an annual index of the 

sum of weighted hourly concentrations, cumulated over 12 hours per day (8 am to 8 pm), during 

the consecutive 3-month period within the O3 season, with a maximum index value set at a level 

within the range of 7 to 15 ppm-hours. The Agency also solicited CASAC review of the 

proposed rule on January 25, 2010 and solicited additional CASAC advice on January 26, 2011. 

After considering comments from CASAC and the public, the EPA prepared a draft final rule, 

which was submitted for interagency review pursuant to Executive Order 12866. On September 

2, 2011, consistent with the direction of the President, the Administrator of the Office of 

Information and Regulatory Affairs (“OIRA”), Office of Management and Budget (“OMB”), 

returned the draft final rule to the EPA for further consideration. In view of this return and the 

timing of the Agency’s ongoing periodic review of the O3 NAAQS required under Clean Air Act 

section 109 (as announced on September 29, 2008), the EPA decided to coordinate further 

proceedings on its voluntary rulemaking on reconsideration with that ongoing periodic review, 

by deferring the completion of its voluntary rulemaking on reconsideration until it completes its 

statutorily-required periodic review. 

In light of EPA’s decision to consolidate the reconsideration with the current review, the 

Court proceeded with the litigation on the 2008 final decision. On July 23, 2013, the D.C. Circuit 

Court of Appeals upheld EPA’s 2008 primary O3 standard, but remanded the 2008 secondary 

standard to the EPA. State of Mississippi v. EPA, 744 F. 3d 1334. With respect to the primary 

standard, the court first held that the EPA reasonably determined that the existing standard was 

not requisite to protect public health with an adequate margin of safety, and consequently 

required revision. Specifically, the court noted that there were “numerous epidemiological 
1-10 

studies linking health effects to exposure to ozone levels below 0.08 ppm and clinical human 

exposure studies finding a causal relationship between health effects and exposure to ozone 

levels at and below 0.08 ppm”. 744 F. 3d at 1345. The court also specifically endorsed the 

weight of evidence approach utilized by EPA in its deliberations. Id. at 1344. 

The court went on to reject arguments that EPA should have adopted a more stringent 

primary standard. Dismissing arguments that a clinical study (as properly interpreted by EPA) to 

show effects at 0.06 ppm necessitated a standard level lower than that selected, the court noted 

that this was a single, limited study. Id. at 1350. With respect to the epidemiologic evidence, the 

court accepted EPA’s argument that there could be legitimate uncertainty that a causal 

relationship between O3 and 8-hour exposures less than 0.075 ppm exists, so that associations at 

lower levels reported in epidemiologic studies did not necessitate a more stringent standard. Id. 

at 1351-52.10

The court also rejected arguments that an 8-hour primary standard of 0.075 ppm failed to 

provide an adequate margin of safety, noting that margin of safety considerations involved policy 

judgments by the agency, and that by setting a standard “appreciably below” the level of the 

current standard (0.08 ppm), the agency had made a reasonable policy choice. Id. Finally, the 

court rejected arguments that EPA’s decision was inconsistent with CASAC’s scientific 

recommendations because CASAC had been insufficiently clear in its recommendations whether 

it was providing scientific or policy recommendations, and EPA had reasonably addressed 

CASAC’s policy recommendations. Id. at 1357-58. 

With respect to the secondary standard, the court held that because EPA had failed to 

identify a level of air quality requisite to protect public welfare, EPA’s comparison between the 

primary and secondary standards for determining if requisite protection for public welfare was 

afforded by the primary standard was inherently arbitrary. The court thus rejected EPA’s 

determination that the revised 8-hour primary standard afforded sufficient protection of public 

welfare, and remanded the standard to EPA. Id. at 1360-62. 

1.2.3 Current O3 NAAQS Review 

On September 29, 2008, the EPA announced the initiation of a new periodic review of 

the air quality criteria for O3 and related photochemical oxidants and issued a call for 

information in the Federal Register (73 FR 56581, Sept. 29, 2008). A wide range of external 

 

10 The court cautioned, however, that “perhaps more [clinical] studies like the Adams studies will 

yet reveal that the 0.060 ppm level produces significant adverse decrements that simply cannot 

be attributed to normal variation in lung function”, and further cautioned that “agencies may not 

merely recite the terms ‘substantial uncertainty’ as a justification for their actions’”. Id. at 1350, 

1357 (internal citations omitted). 
1-11 

experts, as well as EPA staff, representing a variety of areas of expertise (e.g., epidemiology, 

human and animal toxicology, statistics, risk/exposure analysis, atmospheric science, ecology, 

biology, plant science, ecosystem services) participated in a workshop. This workshop was held 

on October 28-29, 2008 in Research Triangle Park, NC. The workshop provided an opportunity 

for a public discussion of the key policy-relevant issues around which the EPA would structure 

this O3 NAAQS review and the most meaningful new science that would be available to inform 

our understanding of these issues. 

Based in part on the workshop discussions, the EPA developed a draft Integrated Review 

Plan outlining the schedule, process, and key policy-relevant questions that would guide the 

evaluation of the air quality criteria for O3 and the review of the primary and secondary O3

NAAQS. A draft of the IRP was released for public review and comment in September 2009. 

This IRP was the subject of a consultation with the CASAC on November 13, 2009 (74 FR 

54562; October 22, 2009).11 We considered comments received from that consultation and from 

the public in finalizing the plan and in beginning the review of the air quality criteria. The EPA’s 

overall plan and schedule for this review is presented in the Integrated Review Plan for the 

Ozone National Ambient Air Quality Standards.

 12 

As part of the process of preparing the O3 ISA, NCEA hosted a peer review workshop in 

October 29-30, 2008 (73 FR 56581, September 29, 2008) on preliminary drafts of key ISA 

chapters. The CASAC and the public reviewed the first external review draft ISA (U.S. EPA, 

2011b; 76 FR 10893, February 28, 2011) at a meeting held in May 19-20, 2011 (76 FR 23809; 

April 28, 2011). Based on CASAC and public comments, NCEA prepared a second draft ISA 

(U.S. EPA, 2011c; 76 FR 60820, September 30, 2011). CASAC and the public reviewed this 

draft at a January 9-10, 2012 (76 FR 236, December 8, 2011) meeting. Based on CASAC and 

public comments, NCEA prepared a third draft ISA (U.S. EPA 2012a; 77 FR 36534; June 19, 

2012), which was reviewed at a CASAC meeting in September 2012. The EPA released the final 

ISA in February 2013. 

The EPA presented its plans for conducting the Risk and Exposure Assessments (REAs) 

that build on the scientific evidence presented in the ISA, in two planning documents titled 

Ozone National Ambient Air Quality Standards: Scope and Methods Plan for Health Risk and 

Exposure Assessment and Ozone National Ambient Air Quality Standards: Scope and Methods 

 

11 See http://yosemite.epa.gov/sab/sabproduct.nsf/WebProjectsbyTopicCASAC!OpenView for more information on 

CASAC activities related to the current O3 NAAQS review. 

12 EPA 452/R-11-006; April 2011; Available: 

http://www.epa.gov/ttn/naaqs/standards/ozone/data/2011_04_OzoneIRP.pdf 
1-12 

Plan for Welfare Risk and Exposure Assessment (henceforth, Scope and Methods Plans).13 

These planning documents outlined the scope and approaches that staff planned to use in 

conducting quantitative assessments, as well as key issues that would be addressed as part of the 

assessments. We released these documents for public comment in April 2011, and consulted with 

CASAC on May 19-20, 2011 (76 FR 23809; April 28, 2011). In designing and conducting the 

initial health risk and welfare risk assessments, we considered CASAC comments (Samet 2011) 

on the Scope and Methods Plans and also considered public comments. In May 2012, we issued 

a memo titled Updates to Information Presented in the Scope and Methods Plans for the Ozone 

NAAQS Health and Welfare Risk and Exposure Assessments that described changes to elements 

of the scope and methods plans and provided a brief explanation of each change and the reason 

for it. 

In July 2012, EPA made the first drafts of the Health and Welfare REAs available for 

CASAC review and public comment (77 FR 42495, July 19, 2012). The first draft PA was made 

available for CASAC review and public comment in August 2012. These documents were 

reviewed by the CASAC O3 Panel at a public meeting in September 2012. The second draft 

REAs and PA were prepared by EPA in consideration of CASAC (Frey and Samet, 2012a, 

2012b) and public comment and were reviewed by the CASAC O3 Panel at a public meeting on 

March 25-27, 2014. This final PA reflects staff’s consideration of the comments and 

recommendations made by CASAC, and comments made by members of the public, in their 

review of draft versions of the PA. 

1.3 GENERAL APPROACH FOR REVIEW OF THE STANDARDS 

As described in section 1.1 above, this PA presents a transparent evaluation and staff 

conclusions regarding policy considerations related to reaching judgments about the adequacy of 

the current standards and the revisions that are appropriate to consider. Staff considerations and 

conclusions in this document are based on the available body of scientific evidence assessed in 

the ISA (U.S. EPA, 2013), exposure and risk analyses presented in the REAs (U.S. EPA, 2014a, 

b), advice and recommendations from CASAC on the first and second draft REAs and PA and 

other draft and final EPA documents in this review, as well as on public comments. This 

evaluation and associated conclusions on the range of policy options that, in staff’s view, are 

supported by the available scientific evidence and exposure/risk information will inform the 

Administrator’s decisions as to whether the existing primary and/or secondary O3 standards 

should be revised and, if so, what revised standard or standards is/are appropriate. 

 

13 EPA-452/P-11-001 and -002; April 2011; Available: 

http://www.epa.gov/ttn/naaqs/standards/ozone/s_o3_2008_pd.html 
1-13 

Staff’s considerations and conclusions related to the current and alternative primary and 

secondary O3 standards are framed by a series of key policy-relevant questions, expanding upon 

those presented in the IRP at the outset of this review (U.S. EPA, 2011a). Answers to these 

questions in this final PA will inform the Administrator’s decisions as to whether, and if so how, 

to revise the current O3 standards. The first overarching question is as follows. 

 Do the currently available scientific evidence and exposure/risk information, as 

reflected in the ISA and REAs, support or call into question the adequacy of the 

protection afforded by the current O3 standards? 

If the answer to this question, which is informed by staff’s consideration of more specific 

questions related to the primary and secondary standards, suggests that revision of the current 

standards may be appropriate, then staff further considers the currently available evidence and 

information with regard to the following question. 

 What range of potential alternative standards is appropriate to consider based on 

the scientific evidence, air quality analyses, and exposure/risk-based 

information? 

The general approaches for consideration of these overarching questions in review of the primary 

and secondary standards are described separately in sections 1.3.1 and 1.3.2 below. 

1.3.1 Approach for the Primary Standard 

Staff’s approach in this review of the current primary O3 standard takes into 

consideration the approaches used in previous O3 NAAQS reviews. The past and current 

approaches described below are both based, fundamentally, on using EPA’s assessment of the 

current scientific evidence and associated quantitative analyses to inform the Administrator’s 

judgment regarding a primary standard for O3 that is “requisite” (i.e., neither more nor less 

stringent than necessary) to protect public health with an adequate margin of safety. 

In reaching conclusions on options for the Administrator’s consideration, we note that the 

final decision to retain or revise the current primary O3 standard is a public health policy 

judgment to be made by the Administrator. This final decision by the Administrator will draw 

upon the available scientific evidence for O3-attributable health effects, and on analyses of 

population exposures and health risks, including judgments about the appropriate weight to 

assign the range of uncertainties inherent in the evidence and analyses. Our general approach to 

informing these judgments, discussed more fully below, recognizes that the available health 

effects evidence reflects a continuum from relatively higher O3 concentrations, at which 

scientists generally agree that health effects are likely to occur, through lower concentrations, at 

which the likelihood and magnitude of a response become increasingly uncertain. Therefore, in 

developing conclusions in this PA, we are mindful that the Administrator’s ultimate judgments 
1-14 

on the primary standard will most appropriately reflect an interpretation of the available 

scientific evidence and exposure/risk information that neither overstates nor understates the 

strengths and limitations of that evidence and information. This approach is consistent with the 

requirements of sections 108 and 109 of the Act, as well as with how the EPA and the courts 

have historically interpreted the Act. 

Section 1.3.1.1 below provides an overview of the general approach taken in the last 

review of the primary O3 NAAQS (i.e., the 2008 review), and a summary of the rationale for the 

decision on the level of the standard in that review (73 FR 16436). Section 1.3.1.2 presents our 

approach in the current review, including our approach to considering the health evidence and 

exposure/risk information, and considerations regarding ambient O3 concentrations attributable 

to background sources. 

1.3.1.1 Approach Used in the Last Review 

In the 2008 review of the O3 NAAQS, the Administrator considered the available 

scientific evidence and exposure/risk information, the advice and recommendations of CASAC, 

and comments from the public. Based on this, he revised the level of the 8-hour primary O3

standard from 0.08 ppm14 to 0.075 ppm (75 ppb15). In reaching a decision to revise the 1997 8-

hour primary O3 standard, the Administrator noted that much new evidence had become 

available since the 1997 review. He noted that this body of scientific evidence was very robust 

and provided consistent and coherent evidence of an array of O3-related respiratory morbidity 

effects, and possibly cardiovascular-related morbidity, as well as total nonaccidental and 

cardiorespiratory mortality. The Administrator specifically observed that (1) the evidence of a 

range of respiratory-related morbidity effects had been considerably strengthened; (2) newly 

available evidence from controlled human exposure and epidemiologic studies identified people 

with asthma as an important susceptible population for which estimates of respiratory effects in 

the general population likely underestimate the magnitude or importance of these effects; (3) 

newly available evidence about mechanisms of toxicity more completely explained the 

biological plausibility of O3-induced respiratory effects and was beginning to suggest 

mechanisms that may link O3 exposure to cardiovascular effects; and (4) there was relatively 

strong evidence for associations between short-term O3 concentrations and total nonaccidental 

and cardiopulmonary mortality. In the opinion of the Administrator, this very robust body of 

evidence enhanced our understanding of O3- related effects and provided increased confidence 

 

14 Due to rounding convention, the 1997 standard level of 0.08 ppm corresponded to 0.084 ppm (84 ppb). 

15 The level of the O3 standard is specified as 0.075 ppm rather than 75 ppb. However, in this PA we refer to ppb, 

which is most often used in the scientific literature and in the ISA, in order to avoid the confusion that could result 

from switching units when discussing the evidence in relation to the standard level. 
1-15 

that various respiratory morbidity effects and other effects marked by indicators of respiratory 

morbidity are causally related to O3 exposures, and the evidence was highly suggestive that O3 

exposures during the warm O3 season contribute to premature mortality.16 

The Administrator also noted important new health evidence reporting a broad array of 

adverse effects following short-term exposures to O3 concentrations below the level of the 1997 

standard, and concerns for such or related effects in at-risk populations,17 including people with 

asthma or other lung diseases, older adults with increased susceptibility, and those who are likely 

to be vulnerable as a result of spending a lot of time outdoors engaged in physical activity (e.g., 

especially active children and outdoor workers). 

He specifically noted new scientific evidence, which built upon existing evidence, 

demonstrating O3-induced lung function effects and respiratory symptoms in some healthy 

individuals following exposures down to 80 ppb. He also noted very limited new evidence 

demonstrating such effects at exposure concentrations well below 80 ppb. In addition, the 

Administrator noted (1) epidemiologic evidence of statistically significant associations with O3-

related health effects in areas that likely would have met the then-current standard; (2) 

epidemiologic studies conducted in areas that likely would have violated the existing standard 

but which nonetheless reported statistically significant associations that generally extended down 

to ambient O3 concentrations below the level of that standard; (3) the few studies that had 

reported statistically significant associations with respiratory morbidity outcomes and mortality 

in subsets of data that included only days with ambient O3 concentrations below the level of the 

existing standard; and (4) controlled human exposure studies, together with animal toxicological 

studies, that provided considerable support for the biological plausibility of the respiratory 

morbidity associations observed in the epidemiologic studies. Based on the available evidence, 

the Administrator agreed with the CASAC and the majority of public commenters that the 

existing standard was not requisite to protect public health with an adequate margin of safety (73 

FR 16471). 

 

16 73 FR 16470-16471 (March 27, 2008) 

17 Here, as in the ISA, the term “at-risk population” is used to encompass populations or lifestages that have a 

greater likelihood of experiencing health effects related to exposure to an air pollutant due to a variety of factors; 

other terms used in the literature include susceptible, vulnerable, and sensitive. These factors may be intrinsic, such 

as genetic factors, lifestage, or the presence of preexisting diseases, or they may be extrinsic, such as socioeconomic 

status (SES), activity pattern and exercise level, or increased pollutant exposures (U.S. EPA 2013, p. lxx, 8-1, 8-2). 

The courts and the Act’s legislative history refer to these at-risk subpopulations as “susceptible” or “sensitive” 

populations. See, e.g., American Lung Ass’n v. EPA, 134 F. 3d 388, 389 (D.C. Cir. 1998) (“NAAQS must protect 

not only average health individuals, but also ‘sensitive citizens’ – children, for example, or people with asthma, 

emphysema, or other conditions rendering them particularly vulnerable to air pollution” (quoting S. Rep. No. 91-

1196 at 10). 
1-16 

Beyond this focus on the available health evidence, the Administrator also considered 

estimates of O3 exposures and health risks based on analyses where air quality was adjusted to 

simulate just meeting the existing and potential alternative standards. For the various air quality 

simulations, he specifically considered the pattern of estimated reductions in O3 exposures across 

health benchmark concentrations of 80, 70, and 60 ppb. The 80 ppb benchmark reflected an 

exposure concentration for which there was strong evidence for respiratory effects in healthy 

people, including airway inflammation, respiratory symptoms, airway hyperresponsiveness, and 

impaired lung host defense (U.S. EPA, 2007, section 4.7). The 60 ppb benchmark reflected an 

exposure concentration for which the Administrator judged the evidence of such effects to be 

very limited (73 FR 16471). 

The Administrator took note of the magnitudes of estimated health risks for a range of 

health effects, including moderate and large lung function decrements, respiratory symptoms, 

respiratory-related hospital admissions, and nonaccidental and cardiorespiratory mortality. He 

recognized that these quantitative risk estimates for a limited number of specific health effects 

were indicative of a much broader array of O3-related effects, including various indicators of 

morbidity in at-risk populations that we could not analyze in the risk assessment (e.g., school 

absences, increased medication use, emergency department visits). The Administrator concluded 

that quantitative exposure and risk estimates, as well as the broader array of O3-related health 

endpoints that could not be quantified, provided additional support for the evidence-based 

conclusion that the existing standard needed to be revised (73 FR 16472). 

Based on the above considerations, and consistent with CASAC’s unanimous conclusion 

that there was no scientific justification for retaining the existing standard, the Administrator 

concluded that the primary O3 standard set in 1997 was not sufficient and thus not requisite to 

protect public health with an adequate margin of safety. He further concluded that revision of 

this standard was needed to provide increased public health protection (73 FR 16472). 

Throughout the 2008 review, CASAC supported a standard level in the range of 60 to 70 

ppb (without change to the form, indicator, or averaging time). In a letter to the Administrator on 

the second draft Staff Paper, CASAC unanimously recommended “that the current primary 

ozone standard be revised and that the level that should be considered for the revised standard be 

from 0.060 to 0.070 ppm’’ (60 to 70 ppb) (Henderson, 2006, p. 5). This recommendation, based 

in part on the placement of more weight on the evidence for effects following exposures to 60 

ppb O3, followed from the CASAC’s more general recommendation that the 1997 standard 

needed to be made substantially more protective of human health, particularly for at-risk 

populations. In a subsequent letter sent specifically to offer advice to aid the Administrator and 

Agency staff in developing the 2007 O3 proposal, CASAC reiterated that Panel members ‘‘were 
1-17 

unanimous in recommending that the level of the current primary ozone standard should be 

lowered from 0.08 ppm to no greater than 0.070 ppm’’ (Henderson, 2007, p. 2).18

After considering CASACs comments, the Administrator judged that the appropriate 

balance to draw, based on the entire body of evidence and information available in the 2008 

review, was a standard set at a level of 75 ppb (and leaving all other elements of the NAAQS 

unchanged). In making this decision, the Administrator placed primary emphasis on the body of 

available scientific evidence, while viewing the results of exposure and risk assessments as 

providing supporting information. Specifically, the Administrator judged that a standard set at 

75 ppb would be appreciably below 80 ppb, the level in controlled human exposure studies at 

which adverse effects had been demonstrated at the time, and would provide a significant 

increase in protection compared to the then-current standard. Based on results of the exposure 

assessment, he also noted that exposures to O3 concentrations at and above a benchmark level of 

80 ppb would be essentially eliminated with a standard level of 75 ppb, and that exposures at and 

above a 70 ppb benchmark level would be substantially reduced or eliminated for the vast 

majority of people in at-risk groups. In addition, the Administrator concluded that the body of 

evidence did not support setting a lower standard level, specifically judging that the available 

evidence for effects following exposures to O3 concentrations of 60 ppb was “too limited to 

support a primary focus at this level” (75 FR 2938). With respect to the epidemiologic evidence, 

the Administrator stated that a standard set at a level lower than 75 ppb “would only result in 

significant further public health protection if, in fact, there is a continuum of health risks in areas 

with 8-hour average O3 concentrations that are well below the concentrations observed in the key 

controlled human exposure studies and if the reported associations observed in the 

epidemiological studies are, in fact, causally related to O3 at those lower levels” (73 FR 16483). 

In making his final decision about the level of the primary O3 standard, the Administrator 

noted that the level of 75 ppb was above the range recommended by CASAC (i.e., 70 to 60 ppb). 

He concluded that “CASAC’s recommendation appeared to be a mixture of scientific and policy 

considerations” (75 FR 2992). The Administrator reached a different policy judgment than the 

CASAC Panel, placing less weight than CASAC on the available controlled human exposure 

studies reporting effects following exposures to 60 ppb O3 and less weight on the results from 

exposure and risk assessments, particularly on estimates of exposures to O3 concentrations at or 

above 60 ppb (73 FR 16482-3). 

 

18 The D.C. Circuit, in its review of the 2008 primary standard, stated that it was unclear whether CASAC’s advice 

reflected issues of pure science or issues of science and policy. That is, the court was unable to determine whether 

CASAC’s conclusion in its 2007 letter that the standard be set no higher than 70 ppb “was based on its scientific 

judgment that adverse effects would occur at that level or instead based on its more qualitative judgment that the 

range it proposed would be more appropriately protective of human health with an adequate margin of safety.” 

Mississippi. 744 F. 3d at 1357. 
1-18 

1.3.1.2 Approach for the Current Review 

To identify the range of options appropriate for the Administrator to consider in the 

current review, we apply an approach that builds upon the general approach used in the last 

review (and in the 2010 reconsideration proposal) and that reflects the broader body of scientific 

evidence, updated exposure/risk information, and advances in O3 air quality modeling now 

available. As summarized above, the Administrator’s decisions in the prior review were based on 

an integration of information on health effects associated with exposure to O3, judgments on the 

adversity and public health significance of key health effects, and expert and policy judgments as 

to when the standard is requisite to protect public health with an adequate margin of safety. 

Staff’s conclusions on the primary O3 standard reflect our consideration of the available 

scientific evidence, exposure/risk information, and air quality modeling information, within the 

context of the overarching questions related to: (1) the adequacy of the current primary O3

standard to protect against effects associated with both short- and long-term exposures and (2) 

potential alternative standards that are appropriate to consider in this review. In addressing these 

broad questions, we organize the discussions in chapters 3 and 4 of this document around a series 

of more specific questions reflecting different aspects of each overarching question. When 

evaluating the health protection afforded by the current and potential alternative standards, we 

take into account the four basic elements of the NAAQS: the indicator, averaging time, form, 

and level. 

Figure 1-1 below provides an overview of our approach in this review. We believe that 

the general approach summarized in this section, and outlined in Figure 1-1, provides a 

comprehensive basis to help inform the judgments required of the Administrator in reaching 

decisions about the current and potential alternative primary O3 standards. In the subsections 

below, we describe our general approaches to considering the scientific evidence (evidencebased considerations) and to considering the human exposure- and health risk information 

(exposure- and risk-based considerations). We also recognize considerations related to ambient 

O3 attributable to background sources.
1-19 

Evidence-Based Considerations

ISA weight-of-evidence conclusions for health effects and at-risk 

populations

Controlled human exposure and animal toxicology studies: Nature, 

magnitude, likely adversity of effects; consistency across studies?

Epidemiologic studies: Statistical precision; confounding; link effects to O3

air quality in specific locations? 

Health effect associations reported in locations meeting current 

standard? 

Confidence in concentration-response relationships over distributions 

of ambient concentrations, including concentrations below level of 

current standard? 

Uncertainties in evidence for O3-attriubutable effects across distributions 

of exposure concentrations and ambient concentrations 

Adequacy of Current 8-Hour Primary O3 Standard

Consider Potential Alternative Standards

Indicator

Support for retaining O3?

Averaging Time

Support for 8-hour only? 

Support for different 

averaging time?

Level

Evidence-based considerations : Consider controlled human exposure studies and epidemiologic 

studies, including uncertainties 

Exposure- and risk-based considerations : Consider exposure and risk reductions for alternative O3

standards, exposures and risks estimated to remain upon meeting alternatives, and uncertainties and 

limitations in exposure/risk estimates 

Form

Support for retaining annual 4th

highest?

Consider retaining 

current 8-hour O3

standard

Identify range of potential alternative standards for consideration 

YES

NO

Does information call 

into question adequacy 

of current 8-hour 

Primary O3 standard?

Exposure-/Risk-Based 

Considerations

Nature, magnitude, and importance of 

estimated exposures and risks 

associated with current O3 standard 

Focus on at-risk populations

Uncertainties in the exposure and risk 

estimates 

Figure 1-1. Overview of approach to reviewing the primary standard. 
1-20 

Consideration of the Scientific Evidence 

Our approach in this review draws upon an integrative synthesis of the entire body of 

available scientific evidence for O3-related health effects, including the evidence newly available 

in the current review and the evidence from previous reviews, as presented in the ISA (U.S. 

EPA, 2013).19 Our approach to considering the scientific evidence is based fundamentally on 

using information from controlled human exposure and epidemiologic studies, supplemented by 

information from animal toxicology studies. Such evidence informs our consideration of the 

health endpoints and at-risk populations on which to focus the current review, and our 

consideration of the O3 concentrations at which various health effects can occur. 

Since the 2008 review of the O3 NAAQS, the Agency has developed formal frameworks 

for characterizing the strength of the scientific evidence with regard to health effects associated 

with exposures to O3 in ambient air and factors that may increase risk in some populations or 

lifestages (U.S. EPA, 2013, Preamble; Chapter 8). These frameworks provide the basis for 

robust, consistent, and transparent processes for evaluating the scientific evidence, including 

uncertainties in the evidence, and for drawing weight-of-evidence conclusions on air pollutionrelated health effects and at-risk populations. 

With regard to characterization of health effects, the ISA uses a five-level hierarchy to 

classify the overall weight-of-evidence into one of the following categories: causal relationship, 

likely to be a causal relationship, suggestive of a causal relationship, inadequate to infer a causal 

relationship, and not likely to be a causal relationship (U.S. EPA, 2013, Preamble Table II). In 

using the weight of evidence approach to inform judgments about the degree of confidence that 

various health effects are likely to be caused by exposure to O3, confidence increases as the 

number of studies consistently reporting a particular health endpoint grows and as other factors, 

such as biological plausibility and strength, consistency and coherence of evidence increases. 

Conclusions about biological plausibility, consistency and coherence of O3-related health effects 

are drawn from the integration of epidemiologic studies with mechanistic information from 

controlled human exposure and animal toxicological studies, as discussed in the ISA (U.S. EPA, 

2013, EPA Framework for Causal Determination, p. 1viii). In this PA, we place the greatest 

 

19Selection of studies for inclusion in the ISA is based on the general scientific quality of the study, and 

consideration of the extent to which the study is informative and policy-relevant. Policy relevant and 

informative studies include those that provide a basis for or describe the relationship between the criteria 

pollutant and effects. This includes studies that offer innovation in method or design and studies that 

reduce uncertainty on critical issues, such as analyses of confounding or effect modification by 

copollutants or other variables, analyses of concentration-response or dose-response relationships, or 

analyses related to time between exposure and response. Review articles, by contrast, are generally not 

included because they typically present summaries or interpretations of existing studies. The specific 

criteria applied to the various types of studies are discussed in more detail in the Preamble to the ISA 

(U.S. EPA, 2013, Preamble). 
1-21 

weight on the health effects for which the evidence has been judged in the ISA to demonstrate a 

causal or a “likely to be” causal relationship with O3 exposures. Our consideration of the 

available evidence for such effects is presented below in Chapter 3 (consideration of the 

adequacy of the current standard) and in Chapter 4 (consideration of potential alternative 

standards). 

As discussed below, we further consider the evidence base assessed in the ISA with 

regard to the types and levels of exposure at which health effects are indicated. This further 

consideration of the evidence, which directly informs EPA’s conclusions regarding the adequacy 

of current or potential alternative standards in providing requisite public health protection, differs 

from consideration of the evidence in the ISA with regard to overarching determinations of 

causality. Therefore, studies that inform determinations of causality may or may not be 

concluded to be informative with regard to the adequacy of the current or potential alternative 

standards.20 

As with health endpoints, the ISA’s characterization of the weight-of-evidence for 

potential at-risk populations is based on the evaluation and synthesis of evidence from across 

scientific disciplines. The ISA characterizes the evidence for a number of “factors” that have the 

potential to place populations at increased risk for O3-related effects. The categories considered 

in evaluating the evidence for these potential at-risk factors are “adequate evidence,” “suggestive 

evidence,” “inadequate evidence,” and “evidence of no effect.” These categories are discussed 

in more detail in the ISA (U.S. EPA, 2013, chapter 8, Table 8-1). In this PA, we focus our 

consideration of potential at-risk populations on those factors for which the ISA judges there is 

“adequate” evidence (U.S. EPA, 2013, Table 8-6). At-risk populations are discussed in more 

detail in section 3.2.1, below. 

Using the available scientific evidence to inform conclusions on the current and potential 

alternative standards is complicated by the recognition that a population-level threshold has not 

been identified, below which it can be concluded with confidence that O3-attributable effects do 

not occur (U.S. EPA, 2013, section 2.5.4.4). In the absence of a discernible threshold, our 

general approach to considering the available O3 health evidence involves characterizing our 

confidence in the extent to which O3-attributable effects occur, and the extent to which such 

effects are adverse, over the ranges of O3 exposure concentrations evaluated in controlled human 

exposure studies and over the distributions of ambient O3 concentrations in locations where 

epidemiologic studies have been conducted. As noted above, we recognize that the available 

 

20For example, as discussed further in this section and in Chapters 3 and 4 of this PA, we judge that health studies 

evaluating exposure concentrations near or below the level of the current standard and epidemiologic studies 

conducted in locations meeting the current standard are particularly informative when considering the adequacy of 

the public health protection provided by the current standard. 
1-22 

health effects evidence reflects a continuum from relatively high O3 concentrations, at which 

scientists generally agree that adverse health effects are likely to occur, through lower 

concentrations, at which the likelihood and magnitude of a response become increasingly 

uncertain. Aspects of our approach particular to evidence from controlled human exposure and 

epidemiologic studies, respectively, are discussed below. 

Controlled Human Exposure Studies 

Controlled human exposure studies provide direct evidence of relationships between 

pollutant exposures and human health effects (U.S. EPA, 2013, p.lx). Controlled human 

exposure studies provide data with the highest level of confidence since they provide human 

effects data under closely monitored conditions and can provide exposure response relationships. 

Such studies are particularly useful in defining the specific conditions under which pollutant 

exposures can result in health impacts, including the exposure concentrations, durations, and 

ventilation rates under which effects can occur. As discussed in the ISA, controlled human 

exposure studies provide clear and compelling evidence for an array of human health effects that 

are directly attributable to acute exposures to O3 per se (i.e., as opposed to O3 and other 

photochemical oxidants, for which O3 is an indicator, or other co-occurring pollutants) (U.S. 

EPA, 2013, Chapter 6). Together with animal toxicological studies, which can provide 

information about more serious health outcomes as well as the effects of long-term exposures 

and mode of action, controlled human exposure studies also help to provide biological 

plausibility for health effects observed in epidemiologic studies. 

In this PA, we consider the evidence from controlled human exposure studies in two 

ways. First, we consider the extent to which controlled human exposure studies provide evidence 

for health effects following exposures to different O3 concentrations, down to the lowestobserved-effects levels in those studies. Second, we use such studies to inform our evaluation of 

the extent to which we have confidence in health effect associations reported in epidemiologic 

studies down through lower ambient O3 concentrations, where the likelihood and magnitude of 

O3-attributable effects become increasingly uncertain. 

We consider the range of O3 exposure concentrations evaluated in controlled human 

exposure studies, including concentrations near or below the level of the current standard. We 

consider both group mean responses, which provide insight into the extent to which observed 

changes are due to O3 exposures rather than to chance alone, and inter-individual variability in 

responses, which provides insight into the fraction of the population that might be affected by 

such O3 exposures (U.S. EPA, 2013, section 6.2.1.1). When considering the relative weight to 

place on various controlled human exposure studies, we consider the exposure conditions 

evaluated (e.g., exercising versus resting, exposure duration); the nature, magnitude, and likely 

adversity of effects over the range of reported O3 exposure concentrations; the statistical 
1-23 

precision of reported effects; and the consistency of results across studies for a given health 

endpoint and exposure concentration. In addition, because controlled human exposure studies 

typically involve healthy individuals and do not evaluate the most sensitive individuals in the 

population (U.S. EPA, 2013, Preamble p. lx), when considering the implications of these studies 

for our evaluation of the current and potential alternative standards, we also consider the extent 

to which reported effects are likely to reflect the magnitude and/or severity of effects in at-risk 

groups. 

Epidemiologic Studies 

We also consider epidemiologic studies of short- and long-term O3 concentrations in 

ambient air. Epidemiologic studies provide information on associations between variability in 

ambient O3 concentrations and variability in various health outcomes, including lung function 

decrements, respiratory symptoms, school absences, hospital admissions, emergency department 

visits, and premature mortality (U.S. EPA, 2013, Chapters 6 and 7). Epidemiologic studies can 

inform our understanding of the effects in the study population (which may include at-risk 

groups) of real-world exposures to the range of O3 concentrations in ambient air, and can provide 

evidence of associations between ambient O3 levels and serious acute and chronic health effects 

that cannot be assessed in controlled human exposure studies. For these studies, the degree of 

uncertainty introduced by confounding variables (e.g., other pollutants, temperature) and other 

factors affects the level of confidence that the health effects being investigated are attributable to 

O3 exposures, alone and in combination with copollutants. 

Available studies have generally not indicated a discernible population threshold, below 

which O3 is no longer associated with health effects (U.S. EPA, 2013, section 2.5.4.4). However, 

the currently available epidemiologic evidence indicates decreased confidence in reported 

concentration-response relationships for O3 concentrations at the lower ends of ambient 

distributions (U.S. EPA, 2013, section 2.5.4.4). Therefore, our general approach to considering 

the results of epidemiologic studies within the context of the current and potential alternative 

standards focuses on characterizing the range of ambient O3 concentrations over which we have 

the most confidence in O3-associated health effects, and the concentrations below which our 

confidence in such health effect associations becomes appreciably lower. In doing so, we 

consider the statistical precision of O3 health effect associations reported in study locations with 

various ambient O3 concentrations; confidence intervals around concentration-response functions 

reported over distributions of ambient O3 (where available); and the extent to which the 

biological plausibility of associations at various ambient O3 concentrations is supported by 

evidence from controlled human exposure and/or animal toxicological studies. 

We consider both multi-city and single-city studies assessed in the ISA, each of which 

have strengths and limitations. Multi-city studies evaluate large populations and provide greater 
1-24 

statistical power than single-city studies. Multi-city studies also reflect O3-associated health 

impacts across a range of diverse locations, providing spatial coverage for different regions and 

reflecting differences in exposure-related factors that could impact O3 risks. In addition, 

compared to single-city studies, multi-city studies are less prone to publication bias and they 

afford the possibility of generalizing to the broader national population (U.S. EPA, 2004, p. 8-

30). In contrast, while single-city studies are more limited than multicity studies in terms of 

statistical power and geographic coverage, conclusions regarding the extent to which air quality 

met the current or potential alternative standards in the cities for which associations have been 

reported can be made with greater certainty for single-city studies (compared to multicity studies 

reporting only multicity effect estimates) because the associations are reported for city-specific 

analyses (U.S. EPA, 2011d, section 2.3.4.1).21 In some cases, single-city studies can also 

provide evidence for locations or population-specific characteristics not reflected in multicity 

studies (U.S. EPA, 2013, section 6.2.7.1). Therefore, when considering available epidemiologic 

studies we evaluate both multi-city and single-city studies, recognizing the strengths and 

limitations of each. 

In placing emphasis on specific epidemiologic studies, we focus on studies conducted in 

the U.S. and Canada. Such studies reflect air quality and exposure patterns that are likely more 

typical of the U.S. population than the air quality and exposure patterns reflected in studies 

conducted outside the U.S. and Canada.22 We also focus on studies reporting associations with 

effects judged in the ISA to be robust to confounding by other factors, including co-occurring air 

pollutants. 

Consideration of Exposure and Risk Estimates 

To put judgments about O3-related health effects into a broader public health context, we 

consider exposure and risk estimates from the HREA, which develops and applies models to 

estimate human exposures to O3 and O3-related health risks in urban case study areas across the 

United States (U.S. EPA, 2014a). The HREA estimates exposures of concern, based on 

interpreting quantitative exposure estimates within the context of controlled human exposure 

study results; lung function risks, based on applying exposure-response relationships from 

controlled human exposure studies to quantitative estimates of exposures; and epidemiologicbased risk estimates, based on applying concentration-response relationships drawn from 

epidemiologic studies to adjusted air quality. Each of these types of assessments is discussed 

briefly below. 

 

21 Though in some cases multicity studies present single-city effect estimates in addition to multi-city estimates. 

22 Though we recognize that a broader body of studies, including international studies, inform the causal 

determinations in the ISA. 
1-25 

As in the 2008 review, the HREA estimates exposures at or above benchmark 

concentrations of 60, 70, and 80 ppb, reflecting exposure concentrations of concern based on the 

available health evidence.23 Estimates of exposures of concern, defined as personal exposures 

while at moderate or greater exertion to 8-hour average ambient O3 levels, at or above these 

discrete benchmark concentrations provide perspective on the public health impacts of O3-related 

health effects that are plausibly linked to the more serious effects seen in epidemiological 

studies, but cannot be evaluated in quantitative risk assessments. They also help elucidate the 

extent to which such impacts may be reduced by meeting the current and alternative standards. 

Estimates of the number of people likely to experience exposures of concern cannot be directly 

translated into quantitative estimates of the number of people likely to experience specific health 

effects due to individual variability in responsiveness. Only a subset of individuals can be 

expected to experience such adverse health effects, and at-risk populations or lifestages, such as 

people with asthma or children, are expected to be affected more by such exposures than healthy 

adults. Though this analysis is conducted using discrete benchmark concentrations, healthrelevant exposures are more appropriately viewed as a continuum with greater confidence and 

less uncertainty about the existence of health effects at higher O3 exposure concentrations and 

less confidence and greater uncertainty at lower exposure concentrations. This approach 

recognizes that there is no sharp breakpoint within the exposure-response relationship for 

exposure concentrations at and above 80 ppb down to 60 ppb. 

The HREA also generates quantitative estimates of O3 health risks for air quality adjusted 

to just meet the current and potential alternative standards. As noted above, one approach to 

estimating O3 health risks is to combine modeled exposure estimates with exposure-response 

relationships derived from controlled human exposure studies of O3-induced health effects. The 

HREA uses this approach to estimate the occurrence of O3-induced lung function decrements in 

at-risk populations in urban case study areas, including school-age children, school-age children 

with asthma, adults with asthma, and older adults. The available exposure-response information 

does not support this approach for other endpoints evaluated in controlled human exposure 

studies (U.S. EPA, 2014a, section 2.2.5 to 2.2.7). 

Another approach to estimating O3-associated health risks is to apply concentrationresponse relationships derived from short- and/or long-term epidemiologic studies to air quality 

adjusted to just meet current and potential alternative standards. The concentration-response 

relationships drawn from epidemiologic studies are based on population exposure surrogates, 

such as 8-hour concentrations averaged across monitors and over more than one day 

 

23 For example, see 75 FR 2945-2946 (January 19, 2010) and 73 FR 16441-16442 (March 27, 2008) discussing 

“exposures of concern”. 
1-26 

(incorporation of lag) (U.S. EPA, 2013, Chapter 6). The HREA presents epidemiologic-based 

risk estimates for O3-associated mortality, hospital admissions, emergency department visits, and 

respiratory symptoms (U.S. EPA, 2014a, Chapter 7). These estimates are derived from the full 

distributions of ambient O3 concentrations estimated for the study locations.24 In addition, the 

HREA estimates mortality risks associated with various portions of distributions of short-term O3

concentrations (U.S. EPA, 2014a). In this PA we consider risk estimates based on the full 

distributions of ambient O3 concentrations and, when available, estimates of the risk associated 

with various portions of those ambient distributions.25 In doing so, we take note of the ISA 

conclusions regarding confidence in linear concentration-response relationships over 

distributions of ambient concentrations, and of the extent to which health effect associations at 

various ambient O3 concentrations are supported by the evidence from experimental studies for 

effects following specific O3 exposures. 

 Considerations Regarding Ambient O3 Concentration Estimates 

Attributable to Background Sources 

As noted above, our approach in this review utilizes recent advances in modeling 

techniques to estimate the contributions of U.S. anthropogenic, international anthropogenic, and 

natural sources to ambient O3 (discussed in detail in Chapter 2 of this PA). Such model estimates 

can provide insights into the extent to which different types of background emissions sources 

contribute to total ambient O3 concentrations. Consideration of this issue in the current review is 

informed by the approaches taken in previous reviews, as well as by court decisions in 

subsequent litigation. 

In 1979, the EPA set a 1-hour O3 standard with a level of 0.12 ppm. Following the final 

decision in that review, the City of Houston argued that the standard was arbitrary and capricious 

because natural O3 concentrations and other physical phenomena in the Houston area made the 

standard unattainable in that area. The D.C. Circuit rejected this argument, stating that 

attainability and technological feasibility are not relevant considerations in the promulgation of 

the NAAQS. The Court also noted that the EPA need not tailor the NAAQS to fit each region or 

locale, pointing out that Congress was aware of the difficulty in meeting standards in some 

 

24 In previous reviews, including the 2008 review and reconsideration, such risks were separately estimated for O3

concentrations characterized as above policy-relevant background concentrations. Policy-relevant background 

concentrations were defined as the distribution of ozone concentrations attributable to sources other than 

anthropogenic emissions of ozone precursor emissions (e.g., VOC, CO, NOx) in the U.S., Canada, and Mexico. The 

decision to estimate total risk across the full range of O3 concentrations reflects current OAQPS views and 

consideration of advice from CASAC (Frey and Samet, 2012b). 

25In a series of sensitivity analyses, the HREA also evaluates a series of threshold models for respiratory mortality 

associated with long-term O3 concentrations. In this PA we consider these risk estimates based on threshold models, 

in addition to HREA core estimates based on the linear model (sections 3.2.3.2, 4.4.2.3). 
1-27 

locations and had addressed this difficulty through various compliance related provisions in the 

Act. See API v. Costle, 665 F.2d 1176, 1184-6 (D.C. Cir. 1981). 

More recently, in the 1997 review of the O3 NAAQS, the Administrator set an 8-hour 

standard with a level of 0.08 ppm (84 ppb). In reaching this decision, the EPA identified several 

reasons supporting its decision to reject a more stringent standard of 0.07 ppm. Most 

importantly, the EPA pointed out the scientific uncertainty at lower concentrations and placed 

significant weight on the fact that no CASAC panel member supported a standard level set lower 

than 0.08 ppm (62 FR 38868). In addition to noting the uncertainties in the health evidence for 

exposure concentrations below 0.08 ppm and the advice of CASAC, the EPA noted that a 

standard set at a level of 0.07 ppm would be closer to peak background concentrations that 

infrequently occur in some areas due to nonanthropogenic sources of O3 precursors (62 FR 

38856, 38868; July 18, 1997). 

In subsequent litigation, the D.C. Circuit upheld the EPA’s decision as the product of 

reasoned decision-making. The Court made clear that the most important support for the EPA’s 

decision was the health evidence and the concerns it raised about setting a standard level below 

0.08 ppm. The Court also pointed to the significant weight that the EPA properly placed on the 

advice it received from CASAC. Finally (as discussed in section 1.2.2 above), the Court noted 

that the EPA could also consider relative proximity to peak natural background O3 when 

considering alternatives within the range of reasonable values supported by the scientific 

evidence and judgments of the Administrator. See ATA III, 283 F.3d at 379 (D.C. Cir. 2002). 

These cases provide a framework for considering the contributions of U.S. 

anthropogenic, international anthropogenic, and natural sources within the context of considering 

the health evidence and CASAC advice, when evaluating various potential alternative standards. 

Consistent with such a framework, this PA identifies the range of policy options for the primary 

O3 standard that staff concludes are appropriate to consider in light of the available scientific 

evidence and exposure/risk information, and the advice of CASAC. In identifying the range of 

policy options supported by the evidence and information, staff has not considered proximity to 

background O3 concentrations. The Administrator, when evaluating the range of possible 

standards that are supported by the scientific evidence, could consider proximity to background 

O3 concentrations as one factor in selecting the appropriate standard.

1.3.2 Approach for the Secondary Standard 

Staff’s approach in this review of the current secondary standard takes into consideration 

aspects of the approaches used in past O3 NAAQS reviews. The past and current approaches, 

generally described below, are both based fundamentally on using EPA’s assessment of the 

current scientific evidence and associated quantitative analyses to inform the Administrator’s 
1-28 

judgment regarding a secondary standard for O3 that is requisite (i.e., neither more nor less 

stringent than necessary) to protect public welfare. 

In reaching conclusions on options for the Administrator’s consideration, we note that the 

final decision to retain or revise the current secondary O3 standard is a public welfare policy 

judgment to be made by the Administrator. This final decision will draw upon the available 

scientific evidence for O3-attributable welfare effects and on analyses of vegetation and 

ecosystem exposures and public welfare risks based on impacts to vegetation, ecosystems and 

their associated services, including judgments about the appropriate weight to place on the range 

of uncertainties inherent in the evidence and analyses. In determining the requisite level of 

protection for crops and trees, the Administrator will need to weigh the importance of the 

predicted risks of these effects in the overall context of public welfare protection, along with a 

determination as to the appropriate weight to place on the associated uncertainties and limitations 

of this information. Our general approach to informing these judgments, discussed more fully 

below, recognizes that the available evidence demonstrates a range of O3 sensitivity across 

studied plant species and documents an array of O3-induced effects that extend from lower to 

higher levels of biological organization. These effects range from those affecting cell processes 

and individual plant leaves to effects on the physiology of whole plants, species effects and 

effects on plant communities to effects on related ecosystem processes and services. Given this 

evidence, it is not possible to generalize across all studied species regarding which cumulative 

exposures are of greatest concern, as this can vary by situation due to differences in exposed 

species sensitivity, the importance of the observed or predicted O3-induced effect, the role that 

the species plays in the ecosystem, the intended use of the affected species and its associated 

ecosystem and services, the presence of other co-occurring predisposing or mitigating factors, 

and associated uncertainties and limitations. At the same time, the evidence also demonstrates 

that though effects of concern can occur at very low exposures in sensitive species, at higher 

cumulative exposures those effects would likely occur at a greater magnitude and/or higher 

levels of biological organization and additional species would likely be impacted. It is important 

to note, however, due to the variability in the importance of the associated ecosystem services 

provided by different species at different exposures and in different locations, as well as 

differences in associated uncertainties and limitations, adverse effects observed or predicted at 

lower exposures along the exposure continuum may or may not have less public welfare 

significance than those observed at higher cumulative exposures. Therefore, in developing 

conclusions in this final PA, we take note of the complexity of judgments to be made by the 

Administrator regarding the adversity of known and anticipated effects to the public welfare and 

are mindful that the Administrator’s ultimate judgments on the secondary standard will most 

appropriately reflect an interpretation of the available scientific evidence and exposure/risk 
1-29 

information that neither overstates nor understates the strengths and limitations of that evidence 

and information. 

Section 1.3.2.1 below provides an overview of the general approach taken in the last 

review of the secondary standard for O3 (i.e., the 2008 review), and a summary of the rationale 

for the decision on the standard in that review (73 FR 16436). Section 1.3.2.2 presents our 

approach in the current review, including our approach to considering the vegetation effects 

evidence and exposure/risk information, and considerations regarding ambient O3 concentrations 

attributable to background sources. 

1.3.2.1 Approach Used in the Last Review 

In the 2008 review of the secondary NAAQS for O3, the Administrator relied upon 

consideration of the available scientific evidence and exposure/risk information, information 

regarding biologically-relevant exposure indices, air quality information regarding the degree of 

overlap between different exposure index forms, the advice and recommendations of CASAC, 

considerations regarding adversity, and comments from the public. Based on all of this, he 

revised the level of the secondary O3 standard from 0.08 ppm26 to 0.075 ppm (75 ppb27). 

In reaching a decision to revise the 1997 8-hour secondary standard, the Administrator 

found, after carefully considering the public comments, that the fundamental scientific 

conclusions on the effects of O3 on vegetation and sensitive ecosystems reached in the 2006 

Criteria Document and 2007 Staff Paper, as discussed in section IV.A of the final rule remained 

valid (73 FR 16496). He further recognized that several additional lines of evidence had 

progressed sufficiently since the 1997 review to provide a more complete and coherent picture of 

the scope of O3-related vegetation risks (i.e., visible foliar injury, tree biomass loss, crop yield 

loss, and others), especially those faced by sensitive seedling, sapling and mature growth stage 

tree species growing in field settings, and their associated forested ecosystems. This new 

research reflected an increased emphasis on field-based exposure methods (e.g., free-air, ambient 

gradient and biomonitoring surveys) (73 FR 16490) in addition to the more traditional controlled 

open-top chamber (OTC) studies (73 FR 16485), and began to address one of the key data gaps 

cited by the Administrator in the 1997 review (73 FR 16486). Specifically, by providing 

additional evidence that O3-induced crop yield loss and tree seedling biomass loss effects 

observed in chambers also occurs in the field, this new research qualitatively increased support 

 

26 As noted earlier, due to rounding convention, the 1997 standard level of 0.08 ppm corresponded to 0.084 ppm (84 

ppb). 

27 As explained above, the level of the O3 standard is specified as 0.075 ppm rather than 75 ppb. However, in this 

draft PA we refer to ppb, which is most often used in the scientific literature and in the ISA, in order to avoid the 

confusion that could result from switching units when discussing the evidence in relation to the standard level. 
1-30 

for, and confidence in, the continued use of OTC-derived crop and tree seedling exposureresponse (E-R) functions developed in the National Crop Loss Assessment Network (NCLAN) 

and National Health and Environmental Effects Research Laboratory – Western Ecology 

Division (NHEERL-WED) studies, respectively, to predict O3-induced impacts on crops and tree 

seedlings in the field (72 FR 37886). All of these areas were considered together, along with 

associated uncertainties, in an integrated weight-of-evidence approach (73 FR 16490). 

Beyond the available vegetation effects evidence, the Administrator also considered 

estimates of O3 exposures and risks when air quality was adjusted to simulate just meeting the 

existing and potential alternative standards. On the basis of these assessments, the Administrator 

concluded that O3 exposures that would be expected to remain after meeting the existing 

standard would be sufficient to cause visible foliar injury and seedling and mature tree biomass 

loss in O3-sensitive vegetation (73 FR 16496) and would still allow O3-related yield loss to occur 

in some commodity crop species and fruit and vegetable species grown in the U.S. (73 FR 

16489). Other O3-induced effects described in the literature, including an impaired ability of 

many sensitive species and genotypes within species to adapt to or withstand other 

environmental stresses, such as freezing temperatures, pest infestations and/or disease, and to 

compete for available resources, would also be anticipated to occur. In the long run, the result of 

these impairments (e.g., loss in vigor) could lead to premature plant death in O3 sensitive species. 

Though effects on other ecosystem components had only been examined in isolated cases, the 

Administrator noted effects such as those described above could have significant implications for 

plant community and associated species biodiversity and the structure and function of whole 

ecosystems (73 FR 16496). 

Although the Administrator concluded that the then-current standard was not sufficient to 

protect against the known and anticipated effects described above, he also recognized that the 

secondary standard is not meant to protect against all known observed or anticipated O3-related 

effects, but only those that can reasonably be judged to be adverse to the public welfare. The 

Administrator found that the degree to which such effects should be considered to be adverse 

depended on the intended use of the vegetation and its significance to the public welfare (73 FR 

16496). In this regard, he took note of a number of actions taken by Congress to establish public 

lands that are set aside for specific uses that are intended to provide benefits to the public 

welfare, including lands that are to be protected so as to conserve the scenic value and the natural 

vegetation and wildlife within such areas, and to leave them unimpaired for the enjoyment of 

future generations. Based on these considerations, and taking into consideration the advice and 

recommendations of CASAC, the Administrator concluded that the protection afforded by the 

existing standard was not sufficient, and that the standard needed to be revised to provide 
1-31 

additional protection from known and anticipated adverse effects on sensitive natural vegetation 

and ecosystems (73 FR 16497). 

Given this judgment, the Administrator then considered what revisions to the standard 

were requisite to protect public welfare. Regarding the form of the standard, the Administrator 

took note that at the conclusion of the 1997 review, the biological basis for a cumulative, 

seasonal form was not in dispute28 and that the 2006 Criteria Document also concluded that O3

exposure indices that cumulate differentially-weighted hourly concentrations are the best 

candidates for relating exposure to plant growth responses (EPA, 2006) (61 FR 65716; 73 FR 

16486). The CASAC, in its letter to the Administrator following its review of the second draft 

Staff Paper, stated that “there is a clear need for a secondary standard which is distinctly 

different from the primary standard in averaging time, level and form” and that “the CASAC 

unanimously agrees that it is not appropriate to try to protect vegetation from the substantial, 

known or anticipated, direct and/or indirect, adverse effects of ambient ozone by continuing to 

promulgate identical primary and secondary standards for ozone” (Henderson, 2006, pp. 5-7). 

Although many possible cumulative, seasonal concentration-weighted exposure metrics exist, the 

Staff Paper and the CASAC Panel concluded that the W12629 form is the most biologicallyrelevant cumulative, seasonal form appropriate to consider in the context of the secondary 

standard review (73 FR 16486-87).30 

Although agreeing with the Criteria Document, Staff Paper and CASAC conclusions that 

a cumulative exposure index that differentially weights O3 concentrations could represent a 

reasonable policy choice for a seasonal secondary standard to protect against the effects of O3 on 

vegetation and that the most appropriate cumulative, concentration-weighted form to consider 

was the sigmoidally-weighted W126 form (73 FR 16498), the Administrator also took note of 

the 1997 decision to make the revised secondary standard identical to a revised primary standard 

after similar considerations (73 FR 16498). In considering the rationale for the 1997 decision, the 

Administrator observed that it was based in part on an analysis that compared the degree of 

 

28 In the 1997 review, a different cumulative metric (SUM06) was proposed. 

29 W126 is a cumulative exposure index that is biologically based. The W126 index focuses on the higher hourly 

average concentrations, while retaining the mid-and lower-level values. It is defined as the sum of sigmoidallyweighted hourly O3 concentrations over a specified period, where the daily sigmoidal weighting function is defined 

as: 1 – exp[-(W126/η)

β

]

30 In a subsequent letter offering unsolicited advice to the Administrator and Agency staff on development of the 

proposed rulemaking, the CASAC reiterated that Panel members ‘‘were unanimous in supporting the 

recommendation in the Final Ozone Staff Paper that protection of managed agricultural crops and natural terrestrial 

ecosystems requires a secondary Ozone NAAQS that is substantially different from the primary ozone standard in 

averaging time, level and form”...and “[t]he recommended metric for the secondary ozone standard is the 

(sigmoidally-weighted) W126 index, accumulated over at least the 12 ‘daylight’ hours and over at least the three 

maximum ozone months of the summer ‘growing season” (Henderson, March 26, 2007, p.3). 
1-32 

overlap in county-level air quality measured in terms of alternative standard forms (62 FR 

38876). Recognizing that significant uncertainty remained in 1997 regarding conclusions drawn 

from such analyses, the Administrator also considered the results of a similar analysis of recent 

monitoring data undertaken in the 2007 Staff Paper to assess the degree of overlap expected 

between the existing standard (4th high, daily maximum 8-hour concentration averaged over 

three years) and potential alternative standards based on W126 cumulative seasonal forms. 

The Administrator noted that this analysis showed significant overlap between the 8-hour 

secondary standard and selected levels of W126 standard forms, with the degree of overlap 

between these potential alternative standards depending greatly on the W126 level selected and 

the distribution of hourly O3 concentrations within the annual and/or 3-year average period. 

From this analysis, the Administrator recognized that a secondary standard set identical to a 

revised primary standard would provide a significant degree of additional protection for 

vegetation as compared to that provided by the existing secondary standard. In further 

considering the significant uncertainties in the available body of evidence and in the exposure 

and risk analyses, and the difficulty in determining at what point various types of vegetation 

effects become adverse for sensitive vegetation and ecosystems, the Administrator focused his 

consideration on a level for an alternative W126 standard (with an annual form) at the upper end 

of the proposed range (i.e., 21 ppm-hours). The Staff Paper analysis showed that at a W126 level 

of 21 ppm-hours, there would be essentially no counties with air quality expected both to exceed 

such an alternative W126 standard and to meet the revised 8-hour primary standard—that is, 

based on this analysis of counties with ambient O3 monitors, a W126-based level of 21 ppmhours would be unlikely to provide additional protection in any areas beyond that likely to be 

provided by the revised 2008 primary standard (73 FR 16499/500). 

The Administrator also considered the Staff Paper finding that the degree of overlap 

between counties (with areas of concern for vegetation) expected to meet an 8-hour level for the 

form of the existing standard and potential alternative levels of a W126-based standard was 

inconsistent across years analyzed. This variation depended greatly on levels selected for a 

W126-based standard and a 3-year average 4th high daily maximum 8-hour standard, 

respectively, and the distribution of hourly O3 concentrations within the annual and/or 3-year 

average period. From this, the Staff Paper recognized the need for caution in evaluating the 

likely vegetation impacts associated with a given level of air quality expressed in terms of the 

existing 8-hour average standard in the absence of parallel W126 information. In considering 

these findings, the Administrator “recognize[d] that the general lack of rural monitoring data 

made uncertain the degree to which the revised 8-hour standard or an alternative W126 standard 

would be protective, and that there was the potential for not providing the appropriate degree of 

protection for vegetation in areas with air quality distributions that resulted in a high cumulative, 
1-33 

seasonal exposure but did not result in high 8-hour average exposures” (73 FR 16500). With 

regard to the 8-hour standard, he also noted that “[w]hile this potential for under-protection was 

clear, the number and size of areas [then] at issue and the degree of risk [was] hard to determine. 

However, such a standard would also tend to avoid the potential for providing more protection 

than is necessary, a risk that would have arisen from moving to a new form for the secondary 

standard despite the significant uncertainty in determining the degree of risk for any exposure 

level and the appropriate level of protection, as well as uncertainty in predicting exposure and 

risk patterns” (73 FR 16500). 

Thus, although the Administrator agreed with the views and recommendations of 

CASAC that a cumulative, seasonal standard was the most biologically relevant way to relate 

exposure to plant growth response, he also recognized that there remained significant 

uncertainties in determining or quantifying the degree of risk attributable to varying levels of O3

exposure, the degree of protection that any specific cumulative, seasonal standard would 

produce, and the associated potential for error in determining the secondary standard that would 

provide a requisite degree of protection—i.e., sufficient but not more than what is necessary. 

Given these significant uncertainties, the Administrator concluded that establishing a new 

secondary standard with a cumulative, seasonal form, at that time, would have resulted in 

uncertain benefits beyond those afforded by the revised primary standard, and therefore, might 

have been more than necessary to provide the requisite degree of protection (73 FR 16500). 

Based on his consideration of these issues (73 FR 16497), the Administrator judged that the 

appropriate balance to be drawn was to set a secondary standard identical in every way to the 

revised 8-hour primary standard of 0.075 ppm. The Administrator believed that such a standard 

would be sufficient to protect public welfare from known or anticipated adverse effects, and did 

not believe that an alternative cumulative, seasonal standard was needed to provide this degree of 

protection (73 FR 16500). 

As noted above, on July 23, 2013 the D.C. Circuit found this approach to be contrary to 

law because the EPA had failed to identify a level of air quality requisite to protect public 

welfare and, therefore, the EPA’s comparison between the primary and secondary standards for 

determining if requisite protection for public welfare was afforded by the primary standard was 

inherently arbitrary. The court remanded the secondary standard to the EPA for further 

consideration. 744 F. 3d at 1360-62. 

1.3.2.2 Approach for the Current Review 

To identify the range of options appropriate for the Administrator to consider in the 

current review, we apply an approach that builds upon the general approach used in the 2008 

review (and in the 2010 reconsideration proposal), and that reflects the broader body of scientific 
1-34 

evidence, updated exposure/risk information, and advances in O3 air quality modeling now 

available. As summarized above, the Administrator’s decisions in the prior review were based on 

an integration of information on welfare effects associated with exposure to O3, judgments on the 

adversity and public welfare significance of key effects, and, expert and policy judgments as to 

when the standard is requisite to protect public welfare. These considerations were informed by 

air quality and related analyses, quantitative exposure and risk assessments, and qualitative 

assessment of impacts that could not be quantified. In performing the evaluation in this 

document, we are additionally mindful of the recent remand of the secondary standard by the 

D.C. Circuit and our approach in the current review incorporates our response to this remand.

Our approach in this review of the secondary O3 standard also reflects our consideration 

of the available scientific evidence, information on biologically-relevant exposure indices, 

exposure/risk information, and air quality modeling information, within the context of 

overarching questions related to: (1) the adequacy of the current secondary O3 standard to protect 

against effects associated with cumulative, seasonal exposures and (2) potential alternative 

standards, if any, that are appropriate to consider in this review. In addressing these broad 

questions, we have organized the discussions in chapters 5 and 6 of this document around a 

series of more specific questions reflecting different aspects of each overarching question. When 

evaluating the welfare protection afforded by the current or potential alternative standards, we 

take into account the four basic elements of the NAAQS: the indicator, averaging time, form, 

and level. 

Figure 1-2 below provides an overview of our approach in this review. We believe that 

the general approach summarized in this section, and outlined in Figure 1-2, provides a 

comprehensive basis to help inform the judgments required of the Administrator in reaching 

decisions about the current and potential alternative secondary O3 standards. In the subsections 

below, we summarize our general approaches to considering the scientific evidence (evidencebased considerations) and to considering the exposure and risk information (exposure- and riskbased considerations). 
1-35 

Adequacy of Current 8-Hour Secondary O3 Standard

Evidence-Based Considerations

ISA conclusions on causality

Evidence for vegetation effects from cumulative 

exposures allowed by the current standard 

Evidence for vegetation effects in locations that would 

likely have met the current standard 

Consider Potential Alternative Standards

Indicator

Support for retaining O3?

Averaging Time

Support for 12 hour diurnal period?

Support for 3-month seasonal 

period?

Support for annual vs. 3-year period?

Level

Evidence-based considerations: Consider O3 exposure concentrations in OTC and field-based studies, 

including uncertainties 

Exposure and risk considerations: Consider the nature, magnitude, and importance of estimated 

exposures and risks associated with potential alternative O3 standards, O3 air quality projections across 

U.S. and case study locations, including uncertainties 

Form

Support for cumulative, seasonal 

form?

Support for peak-weighted form?

Consider retaining 

current 8-hour O3

standard

Identify range of potential alternative standards for consideration 

YES

NO

Does information 

call into question

the adequacy and 

appropriateness of 

current 8-hour 

secondary O3

standard?

Exposure-/Risk-Based Considerations

Nature, magnitude, and importance of 

estimated exposures and risks associated with 

current O3 standard 

Uncertainties in the exposure and risk 

estimates?

Figure 1-2. Overview of approach to reviewing the secondary standard. 
1-36 

 Consideration of the Scientific Evidence 

Our approach in this review draws upon an integrative synthesis of the entire body of 

available scientific evidence for O3-related welfare effects, including the evidence newly 

available in the current review and the evidence from previous reviews, as presented in the ISA 

(U.S. EPA, 2013). Our approach to considering the scientific evidence for effects on vegetation 

is based fundamentally on using information from controlled chamber studies and field-based 

studies. Such evidence informs our consideration of welfare endpoints and at-risk species and 

ecosystems on which to focus the current review, and our consideration of the ambient O3

conditions under which various welfare effects can occur. 

As in each NAAQS review, we consider the entire body of evidence for the subject 

criteria pollutant. With regard to identification of the welfare effects that could be caused by a 

pollutant, we look to controlled exposure studies using chamber or free air methodologies and 

field-based observational, survey and gradient studies. Evaluating all of the evidence together, 

the ISA makes a determination with regard to the strength of the evidence for a causal 

relationship between the air pollutant and specific welfare effects. These determinations inform 

our identification of welfare effects for which the NAAQS may provide protection. 

Since the 2008 review of the O3 NAAQS, the Agency has developed a formal framework 

for characterizing the strength of the scientific evidence with regard to a causal relationship 

between ambient O3 and welfare effects (U.S. EPA, 2013, Preamble; Chapter 9). This framework 

provides the basis for a robust, consistent, and transparent process for evaluating the scientific 

evidence, including uncertainties in the evidence, and for drawing weight-of-evidence 

conclusions regarding air pollution-related welfare effects. In so doing, the ISA uses a five-level 

hierarchy, classifying the overall weight of evidence into one of the following categories: causal 

relationship, likely to be a causal relationship, suggestive of a causal relationship, inadequate to 

infer a causal relationship, and not likely to be a causal relationship (U.S. EPA, 2013, Preamble 

Table II). In our approach here, we place the greatest weight on the evidence for welfare effects 

that have been judged in the ISA to be caused by, or likely to be caused by, O3 exposures. Our 

consideration of the available evidence for such effects is presented below in Chapter 5 

(consideration of the adequacy of the current standard) and in Chapter 6 (consideration of 

potential alternative standards). 

We further consider the evidence base, as assessed in the ISA, with regard to the types 

and levels of exposure at which welfare effects are indicated. This further consideration of the 

evidence base, which directly informs the EPA’s conclusions regarding the adequacy of current 

or potential alternative standards in providing requisite public welfare protection, differs from 

consideration of the evidence in the ISA with regard to overarching determinations of causality. 
1-37 

Studies that have informed determinations of causality may or may not be concluded to be 

informative with regard to the adequacy of the current or potential alternative standards. 

Our approach in this review, as in past reviews, included recognition that the available 

evidence has not provided identification of a threshold in exposure or ambient O3 concentrations 

below which it can be concluded with confidence that O3-attributable vegetation effects do not 

occur across the broad range of O3-sensitive plant species growing within the U.S. This is due in 

part to the fact that research shows that there is variability in sensitivity between and within 

species and that numerous factors, i.e. chemical, physical, biological, and genetic, can influence 

the direction and magnitude of the studied effect (U.S. EPA, 2013, section 9.4.8). In the absence 

of a discernible threshold, our general approach to considering the available O3 welfare evidence 

involves characterizing our confidence in conclusions regarding O3-attributable vegetation 

effects over the ranges of cumulative seasonal O3 exposure values evaluated in chamber studies 

and in field studies in areas where O3-sensitive vegetation are known to occur, as well as 

characterizing the extent to which these effects can be considered adverse. In addition, because 

O3 can indirectly affect other ecosystem components (such as soils, water, and wildlife, and their 

associated goods and services, through its effects on vegetation) our approach also considers 

those indirect effects for which the ISA concludes, based on multiple lines of evidence, including 

mechanistic and physiological processes, to have a causal or likely to be a causal relationship. 

With respect to ecosystem services for which we may have only limited or qualitative 

information regarding an association with O3 exposures, our approach is to consider their policyrelevance in the context of section 109(b) (2) of the CAA which specifies that secondary 

standards provide requisite protection of “public welfare from any ... known or anticipated 

adverse effects associated with the presence of [the] pollutant in the ambient air”. As noted 

above, our approach to informing these judgments, discussed more fully below, recognizes that 

the available welfare effects evidence demonstrates a wide range in O3 sensitivities across 

studied plant species. As a result, at relatively high cumulative O3 exposures, a greater number 

of plant species will show effects, and the magnitude of the observed effects will be greater, 

particularly on the more sensitive species, while at lower cumulative O3 exposures, fewer species 

will demonstrate effects and the magnitude of those observed effects will be less. 

In this review, the evidence base includes quantitative information across a broad array of 

vegetation effects (e.g., growth impairment during seedlings, saplings and mature tree growth 

stages, visible foliar injury, and yield loss in annual crops) and across a diverse set of exposure 

methods from laboratory and field studies. These methods include the more traditional OTC 

studies, as well as field-based exposure studies. While we consider the full breadth of 

information available, we place greater weight on U.S. studies due to the often species-, site-, 

and climate-specific nature of O3-related vegetation responses. We especially weight those 
1-38 

studies that include O3 exposures that fall within the range of those likely to occur in the ambient 

air. Further, our approach in the context of the quantitative exposure and risk assessments 

(discussed below), places greatest emphasis on studies that have evaluated plant response over 

multiple exposure levels and developed exposure-response relationships that allow the prediction 

(estimation) of plant responses over the range of potential alternative standards being assessed. 

In considering the evidence, we recognize differences across different study types in what 

information they provide. For example, because conditions can be controlled in laboratory 

studies, responses in such studies may be less variable and smaller differences may be easier to 

detect. However, the control conditions may limit the range of responses or incompletely reflect 

pollutant bioavailability, so they may not reflect responses that would occur in the natural 

environment. Alternatively, field data can provide important information for assessments of 

multiple stressors or where site-specific factors significantly influence exposure. They are also 

often useful for analyses of larger geographic scales and higher levels of biological organization. 

However, because most field study conditions cannot be controlled, variability is expected to be 

higher and differences harder to detect. The presence of confounding factors can also make it 

difficult to attribute observed effects to specific stressors. 

In considering information from across multiple lines of evidence, our approach is to first 

integrate the evidence from both controlled and field-based studies and assess the coherence and 

consistency across the available evidence for each effect. We then consider the extent to which 

these identified effects should be considered adverse to the public welfare, relying largely on the 

paradigm used in the 2008 review and 2010 proposed reconsideration (e.g., 75 FR 3006). This 

paradigm recognizes that the significance to the public welfare of O3-induced effects on sensitive 

vegetation growing within the U.S. can vary depending on the nature of the effect, the intended 

use of the sensitive plants or ecosystems, and the types of environments in which the sensitive 

vegetation and ecosystems are located. Accordingly, any given O3-related effect on vegetation 

and ecosystems (e.g., biomass loss, crop yield loss, foliar injury) may be judged to have a 

different degree of impact on the public welfare depending, for example, on whether that effect 

occurs in a Class I area, a city park, or commercial cropland. Our approach takes this variation in 

the significance of O3-related vegetation effects into account in evaluating the currently available 

evidence with regard to the extent to which it calls into question the adequacy of the current 

standard and, as appropriate, indicates potential alternative standards that would be appropriate 

for the Administrator to consider. In the 2010 proposed reconsideration, the Administrator 

proposed to place the highest priority and significance on vegetation and ecosystem effects to 

sensitive species that are known to or are likely to occur in federally protected areas such as 

national parks and other Class I areas, or on lands set aside by States, Tribes and public interest 

groups to provide similar benefits to the public welfare (75 FR 3023/24). Our approach in this 
1-39 

review considers whether newly available information would suggest any evolution to this 

paradigm, in particular in the context of considering associated ecosystem services. 

Finally, our approach continues to give great weight to the scientific evidence available in 

this and previous reviews indicating the relevance of cumulative, seasonal, concentrationweighted exposures in inducing vegetation effects. More specifically, in the 2008 and 2010 

reviews, the EPA concluded and the CASAC agreed that the W126 cumulative exposure metric 

was the most appropriate to use in this review to evaluate both the adequacy of the current 

secondary standard and the appropriateness of any potential revisions. As discussed in chapter 5 

in this PA, the information available in this review continues to support the use of such a metric 

and does not call into question the appropriateness of using the W126 metric in this context. 

Therefore, both the WREA and PA continue to express exposures in terms of the W126 index, 

and continue to consider the important policy implications regarding selection of an appropriate 

exposure index for vegetation. Our approach also places primary emphasis on studies that 

evaluated plant response to exposures that were or can be described using such an index. The 

policy-relevant discussions in chapters 5 and 6 focus on vegetation effects evidence and 

exposure/risk information that can be associated with cumulative, seasonal peak-weighted 

exposures, where possible. Discussions pertaining to the adequacy of the current secondary 

standard will consider what cumulative seasonal exposures would be allowed under air quality 

that would just meet the current standard. 

 Consideration of Exposure and Risk Estimates and Air Quality Analyses 

To put judgments about O3-related vegetation and ecosystem effects and services into a 

broader public welfare context, we consider national scale exposure and risk assessments 

described in the WREA (U.S. EPA, 2014b). We particularly focused on the WREA quantitative 

risks related to three types of vegetation effects: foliar injury, biomass loss, and crop yield loss. 

These risks were assessed in a range of WREA analyses variously involving recent O3

monitoring data and/or national-scale adjusted air quality scenarios for the current secondary 

standard and, in some analyses, for a cumulative, seasonal W126 form at one or more levels (15, 

11 and 7 ppm-hours). Our consideration of these WREA results provide insight into the extent to 

which the current or potential alternative standards would be expected to maintain distributions 

of cumulative, seasonal O3 exposures below those associated with adverse vegetation effects. 

 With regard to quantitative O3 risks related to welfare effects and ecosystem services for 

foliar injury, we consider two main analyses in the WREA: a screening-level assessment of 214 

National Parks and a case study focused on three National Parks. In the screening-level 

assessment, O3 concentrations in national parks are assessed using criteria developed from a U.S. 

Forest Service nationwide dataset on foliar injury, ambient O3 concentrations (in terms of W126 
1-40 

index) and soil moisture (which can influence susceptibility of vegetation to foliar injury). 

Additionally, we consider a case study for three Class I areas (Great Smoky Mountain National 

Park, Rocky Mountain National Park, and Sequoia/Kings Canyon National Park). We consider 

results from this case study for three metrics: 1) percent of vegetation cover affected by foliar 

injury; 2) percent of trails affected by foliar injury; 3) estimates of species specific biomass loss 

within the case study area. We also consider qualitative analyses on ecosystem services effects 

for this endpoint. For example, the WREA uses GIS mapping to illustrate where effects may be 

occurring and relates those areas to national scale statistics for recreational use and data on 

hiking trails, campgrounds and other park amenities that intersect with potentially affected areas. 

These are used to identify impacts on ecosystem services associated with recreation in national 

parks. We additionally consider analyses showing associations between elevated O3

concentrations and increased vulnerability to fire risk regimes, insect attacks and impacts on 

hydrological cycles. 

With regard to risks related to biomass and crop yield loss, we consider WREA results 

based on exposure-response functions for tree and crop species that predict the growth or yield 

response of each species, based on the exposure patterns estimated within its growing region. To 

compare exposure-response across species, genotypes or experiments for which absolute 

response values may vary greatly, the WREA instead uses estimates of relative biomass loss for 

trees or yield loss for crops. The WREA develops such estimates nationally and separately for 

more than 100 federally designated Class I areas. Additionally, we consider WREA-developed 

estimates of associated impacts on the agriculture and forestry sectors quantifying how O3

exposure to vegetation is estimated to affect the provision of timber and crops and carbon 

sequestration. We consider estimates for impacts related to tree biomass loss on ecosystem 

services such as pollution removal, carbon storage and sequestration in five urban case study 

areas. We consider biomass and crop yield loss estimates in light of advice from CASAC, as 

discussed in sections 5.3 and 5.4 below. 

In considering the amount of weight to place on the estimates of exposures and risks at or 

above specific W126 values described in the WREA, our approach: 1) evaluates the weight of 

the scientific evidence concerning vegetation effects associated with those O3 exposures; 2) 

considers the importance, from a public welfare perspective, of the O3-induced effects on 

sensitive vegetation and associated ecosystem services that are known or anticipated to occur as 

a result of exposures at selected W126 values; and, 3) recognizes that predictions of effects 

associated with any given O3 exposure may be mitigated or exacerbated by actual conditions in 

the field (i.e., co-occurring modifying environmental and genetic factors). When considering 

analyses in the WREA that involve discrete exposure levels or varying levels of severity of 

effects, our approach to informing these judgments recognizes that the available welfare effects 
1-41 

evidence demonstrates a wide range in O3 sensitivities across studied plant species. As a result, 

at relatively high cumulative O3 exposures, a greater number of plant species will show effects, 

and the magnitude of the observed effects will be greater, particularly on the more sensitive 

species, while at lower cumulative O3 exposures, fewer species will demonstrate effects and the 

magnitude of those observed effects will be less. We recognize that there is no sharp breakpoint 

along this continuum of effects incidence and severity, ranging from concentrations at and above 

the level of the current secondary standard down to the lowest cumulative, seasonal W126 value 

assessed. In considering these results in this PA, we consider both the potential for welfare 

effects and their severity and our understanding of the likelihood of such effects at different O3

exposures. 

1.3.2.2.1 Considerations Regarding Ambient O3 Concentration Estimates 

Attributable to Background Sources 

As noted above, our approach in this review utilizes recent advances in modeling 

techniques to estimate the contributions of U.S. anthropogenic, international anthropogenic, and 

natural sources to ambient O3 (discussed in detail in Chapter 2 of this document). Such model 

estimates can provide insights into the extent to which different types of emissions sources 

contribute to total ambient O3 concentrations. Our consideration of this issue in the current 

review is informed by the approaches taken in previous reviews, and by court decisions on 

subsequent litigation, as discussed in section 1.3.1.2.3 above. Further, in the 1996 proposal, O3

background concentrations were one of the factors the Administrator considered in selecting the 

SUM06 index as a form for an alternative secondary standard. This and other cumulative 

exposure indices under consideration were judged to be equally capable at estimating exposures 

relevant to vegetation, given the lack of evidence for a discernible threshold for vegetation 

effects in general (U.S. EPA 1996, p. 225), which might have provided a scientific basis for 

selecting among different cumulative exposure indices. At that time, the SUM06 metric was 

selected over the W126 metric because it focused on the policy-relevant (above background) 

portion of the total cumulative seasonal exposures reaching plants (62 FR 38856). At the 

conclusion of that review, the Administrator ultimately chose to set the secondary standard 

identical to the primary standard, including using the 8-hour average instead of a cumulative 

seasonal form (62 FR 38868). In the 2008 review, staff analyses concluded that the W126 index 

was more biologically-relevant based on the available science; staff additionally noted, based on 

then-available estimates of background, that this form was also not likely to be significantly 

impacted by background concentrations given the very low weight assigned to lower O3

concentrations by the W126 index (U.S. EPA, 2007 SP, 7-22; 72 FR 37893). In this review, the 

degree to which the total value of the W126 index could be contributed by background 

concentrations was again assessed. Based on a limited analysis, described in chapter 2 of the 
1-42 

PA, background O3 (BGO3) can comprise a non-negligible portion of W126 across the U.S., 

have the greatest contributions to W126 in the intermountain western U.S., and because of the 

sigmoidal weighting function that emphasizes the background contributions to the highest hourly 

ozone values (when BGO3 contributions are generally lowest), proportionally contribute slightly 

less for the W126 than for seasonal means of maximum daily 8 hour average values. As with the 

primary standard, in identifying the range of policy options supported by the evidence and 

information, staff has not considered proximity to background O3 concentrations. The 

Administrator, when evaluating the range of possible standards that are supported by the 

scientific evidence, could consider proximity to background O3 concentrations as one factor in 

selecting the appropriate standard. 

1.3.3 Organization of this Document 

Chapter 2 of this PA provides an overview of the O3 ambient monitoring network and O3

air quality, including estimates of O3 concentrations attributable to background sources. The 

remaining chapters are organized into two main parts. Chapters 3 and 4 focus on the review of 

the primary O3 NAAQS while chapters 5 and 6 focus on the review of the secondary O3

NAAQS. Staff’s considerations and conclusions related to the current primary and secondary 

standards are discussed in chapters 3 and 5, respectively. Staff’s considerations and conclusions 

related to potential alternative primary and secondary standards are discussed in chapters 4 and 

6, respectively. Key uncertainties in the review and areas for future research and data collection 

are additionally identified in chapters 4 and 6 for the two types of standards. 
1-43 

1.4 REFERENCES 

Frey, H.C. and Samet, J.M. (2012a) CASAC Review of the EPA’s Policy Assessment for the Review of the Ozone 

National Ambient Air Quality Standards (First External Review Draft – August 2012). EPA-CASAC-13-

003. November 26, 2012. 

Frey, H.C. and Samet, J.M. (2012b) CASAC Review of the EPA’s Health Risk and Exposure Assessment for Ozone 

(First External Review Draft - Updated August 2012) and Welfare Risk and Exposure Assessment for 

Ozone (First External Review Draft - Updated August 2012). EPA-CASAC-13-002. November 19, 2012 

Henderson, R. (2006) Letter from CASAC Chairman Rogene Henderson to EPA Administrator Stephen Johnson. 

October 24, 2006, EPA-CASAC-07-001. 

Henderson, R. (2007) Letter from CASAC Chairman Rogene Henderson to EPA Administrator Stephen Johnson. 

March 26, 2007, EPA-CASAC-07-002. 

Henderson, R. (2008) Letter from CASAC Chairman Rogene Henderson to EPA Administrator Stephen Johnson. 

April 7, 2008, EPA-CASAC-08-001. 

Samet, J.M. (2011) Clean Air Scientific Advisory Committee (CASAC) Response to Charge Questions on the 

Reconsideration of the 2008 Ozone National Ambient Air Quality Standards. EPA-CASAC-11-004. 

March 30, 2011. Available online at: 

http://yosemite.epa.gov/sab/sabproduct.nsf/0/F08BEB48C1139E2A8525785E006909AC/$File/EPACASAC-11-004-unsigned+.pdf 

U.S. DHEW (Department of Health, Education, and Welfare) (1970). Air Quality Criteria for Photochemical 

Oxidants. Washington, D.C.: National Air Pollution Control Administration; publication no. AP-63. 

Available from: NTIS, Springfield, VA; PB-190262/BA. 

U.S. EPA (U.S. Environmental Protection Agency). (1978). Air quality criteria for ozone and other photochemical 

oxidants [EPA Report]. (EPA/600/8-78/004). Washington, D.C.. 

U.S. EPA (U.S. Environmental Protection Agency). (1982). Air quality criteria document for ozone and other 

photochemical oxidants. Fed Reg 47: 11561. 

U.S. EPA (U.S. Environmental Protection Agency). (1986). Air quality criteria for ozone and other photochemical 

oxidants [EPA Report]. (EPA-600/8-84-020aF - EPA-600/8-84-020eF). Research Triangle Park, NC. 

http://www.ntis.gov/search/product.aspx?ABBR=PB87142949 

U.S. Environmental Protection Agency. (1989) Review of the National Ambient Air Quality Standards for Ozone: 

Policy Assessment of Scientific and Technical Information. OAQPS Staff Paper. Office of Air Quality 

Planning and Standards, Research Triangle Park, NC. 

U.S. EPA. (1996) Air Quality Criteria for Ozone and Related Photochemical Oxidants Volume I of III (Final, 1996). 

U.S. Environmental Protection Agency, Washington, D.C., EPA/600/AP-93/004aF (NTIS PB94173127). 

U.S. EPA. (2004) Air Quality Criteria for Particulate Matter (Final Report). U.S. Environmental Protection Agency, 

Washington, D.C., EPA 600/P-99/002aF-bF, 2004. 

U.S. Environmental Protection Agency. (2006). Air quality criteria for ozone and related photochemical oxidants 

[EPA Report]. (EPA/600/R-05/004AF). Research Triangle Park, NC. 

http://cfpub.epa.gov/ncea/cfm/recordisplay.cfm?deid=149923 
1-44 

U.S. Environmental Protection Agency. (2007) Review of the National Ambient Air Quality Standards for Ozone: 

Policy Assessment of Scientific and Technical Information. OAQPS Staff Paper. Office of Air Quality 

Planning and Standards, Research Triangle Park, NC. EPA-452/R-07-007. 

U.S. Environmental Protection Agency. (2011a). Integrated Review Plan for the O3 National Ambient Air Quality 

Standards (IRP) U. S. Environmental Protection Agency, National Center for Environmental Assessment 

Office of Research and Development and Office of Air Quality Planning and Standards Office of Air and 

Radiation, Research Triangle Park, North Carolina EPA. 452/R-11-006. 

U.S. Environmental Protection Agency. (2011b) Integrated Science Assessment for Ozone and Related 

Photochemical Oxidants: First External Review Draft, U.S. Environmental Protection Agency, 

Washington, D.C., EPA/600/R-10/076A. 

U.S. Environmental Protection Agency. (2011c) Integrated Science Assessment of Ozone and Related 

Photochemical Oxidants (Second External Review Draft). U.S. Environmental Protection Agency, 

Washington, D.C., EPA/600/R-10/076B. 

U.S. Environmental Protection Agency. (2011d) Policy Assessment for the Review of the Particulate Matter 

National Ambient Air Quality Standards. Office of Air Quality Planning and Standards, Research Triangle 

Park, NC. EPA 452/R-11-003. 

U.S. Environmental Protection Agency. (2012a). Integrated Science Assessment for Ozone and Related 

Photochemical Oxidants: Third External Review Draft, U.S. Environmental Protection Agency, Research 

Triangle Park, NC. EPA/600/R-10/076C 

U.S. Environmental Protection Agency. (2012b). Health Risk and Exposure Assessment for Ozone, First External 

Review Draft, U.S. Environmental Protection Agency, Research Triangle Park, NC. EPA 452/P-12-001. 

U.S. Environmental Protection Agency. (2013). Integrated Science Assessment for Ozone and Related 

Photochemical Oxidants (Final Report). U.S. Environmental Protection Agency, Washington, D.C., 

EPA/600/R-10/076F, 2013 

U.S. Environmental Protection Agency. (2014a). Health Risk and Exposure Assessment for Ozone, Second External 

Review Draft. Office of Air Quality Planning and Standards, Research Triangle Park, NC. EPA-452/P-14-

004a. 

U.S. Environmental Protection Agency. (2014b). Welfare Risk and Exposure Assessment for Ozone, Second 

External Review Draft. Office of Air Quality Planning and Standards, Research Triangle Park, NC. EPA452/P-14-003a. 
2-1 

2 O3 MONITORING AND AIR QUALITY 

 This section provides overviews of ambient O3 monitoring in the U.S. (section 2.1); O3

precursor emissions and atmospheric chemistry (section 2.2); ambient O3 concentrations (section 

2.3); and available evidence and information related to background O3 (section 2.4). These issues 

are also discussed in detail in chapter 3 of the Integrated Science Assessment (ISA) (US EPA, 

2013). 

2.1 O3 MONITORING 

2.1.1 O3 Monitoring Network 

 To monitor compliance with the NAAQS, state and local environmental agencies operate 

O3 monitoring sites at various locations, depending on the population of the area and typical peak 

O3 concentrations.1

 All of the state and local monitoring stations that report data to the EPA AQS 

use ultraviolet (UV) Federal Equivalent Methods (FEMs). The Federal Reference Method (FRM) 

is no longer used due to lack of availability and safety concerns.2

 In 2010, there were over 1,300 

state, local, and tribal O3 monitors reporting concentrations to EPA. The “State and Local 

Monitoring Stations” (SLAMS) minimum monitoring requirements to meet the O3 design criteria 

are specified in 40 CFR Part 58, Appendix D. The requirements are both population and design 

value based.3 The minimum number of O3 monitors required in a Metropolitan Statistical Area 

(MSA) ranges from zero for areas with a population of at least 50,000 and under 350,000 with no 

recent history of an O3 design value greater than 85 percent of the NAAQS, to four for areas with 

a population greater than 10 million and an O3 design value greater than 85 percent of the NAAQS. 

Within an O3 network, at least one site for each MSA, or Combined Statistical Area (CSA) if 

multiple MSAs are involved, must be designed to record the maximum concentration for that 

particular metropolitan area. Since highest O3 concentrations tend to be associated with particular 

seasons for various locations, EPA requires ozone monitoring during specific ozone monitoring 

seasons which vary by state.4

 

1

 The minimum O3 monitoring network requirements for urban areas are listed in Table D-2 of Appendix D to 40 

CFR Part 58. 

2

 EPA is developing a new O3 Federal Reference Method (FRM) and proposed changes to the FEM testing 

requirements to reflect new and improved measurement technology. 

3

 A design value is a statistic that describes the air quality status of a given area relative to the level of the NAAQS. 

Design values are typically used to classify nonattainment areas, assess progress towards meeting the NAAQS, and 

develop control strategies. See http://epa.gov/airtrends/values.html (U, 2010, 677582) for guidance on how these 

values are defined. 

4

 The required O3 monitoring seasons for each state are listed in Table D-3 of Appendix D to 40 CFR Part 58. EPA 

plans to complete an analysis using certified data for the years of 2010-2012 to determine if any changes to the 

length of the required O3 monitoring seasons would be needed to support a revised NAAQS. 
2-2 

Figure 2-1 shows the locations of the U.S. ambient O3 monitoring sites reporting data to 

EPA at any time during the 2006-2010 period. The gray dots which make up over 80% of the O3

monitoring network are SLAMS monitors, which are operated by state and local governments to 

meet regulatory requirements and provide air quality information to public health agencies. Thus, 

the SLAMS monitoring sites are largely focused on urban and suburban areas. The blue dots 

highlight two important subsets of monitoring sites within the SLAMS network: the “National 

Core” (NCore) multi-pollutant monitoring network and the “Photochemical Assessment 

Monitoring Stations” (PAMS) network.5

While the existing U.S. O3 monitoring network has a largely urban focus, to address 

ecosystem impacts of O3 such as biomass loss and foliar injury, it is equally important to focus on 

O3 monitoring in rural areas. The green dots in Figure 2-1 represent the Clean Air Status and 

Trends Network (CASTNET) monitors which are located in rural areas. There were about 80 

CASTNET sites operating in 2010, with sites in the eastern U.S. being operated by EPA and sites 

in the western U.S. being operated by the National Park Service (NPS). Finally, the black dots 

represent “Special Purpose Monitoring Stations” (SPMS), which include about 20 rural monitors 

as part of the “Portable O3 Monitoring System” (POMS) network operated by the NPS. Between 

the CASTNET, NCore, and POMS networks, there were about 120 rural O3 monitoring sites 

operating in the U.S. in 2010. 

 

 

5

 EPA is currently developing proposed revisions to the PAMS network design intended to increase coverage and 

allow for more locale-specific flexibility. 
2-3 

Figure 2-1. Map of U.S. ambient O3 monitoring sites reporting data to EPA during the 

2006-2010 period. 

2.1.2 Recent O3 Monitoring Data and Trends 

To determine whether or not the O3 NAAQS has been met at an ambient monitoring site, 

a statistic commonly referred to as a “design value” must be calculated based on three consecutive 

years of data collected from that site. The form of the existing O3 NAAQS design value statistic is 

the 3-year average of the annual 4th highest daily maximum 8-hour O3 concentration in parts per 

billion (ppb), with decimal digits truncated. The existing primary and secondary O3 NAAQS are 

met at an ambient monitoring site when the design value is less than or equal to 75 ppb.6

 In counties 

or other geographic areas with multiple monitors, the area-wide design value is defined as the 

design value at the highest individual monitoring site, and the area is said to have met the NAAQS 

if all monitors in the area are meeting the NAAQS. 

Figure 2-2 shows the trend in the annual 4th highest daily maximum 8-hour O3

concentrations in ppb based on 933 “trends” sites with complete data records over the 2000 to 

2012 period. The center line in this figure represents the median value across the trends sites, while 

 

6

 For more details on the data handling procedures used to calculate design values for the existing O3 NAAQS, see 

40 CFR Part 50, Appendix P. 
2-4 

the dashed lines represent the 25th and 75th percentiles, and the bottom and top lines represent the 

10th and 90th percentiles. Figures 2-3 and 2-4 show maps of the O3 design values (ppb) at all U.S. 

monitoring sites for the 2009-2011 and 2010-2012 periods, respectively. The trend figure shows 

that the annual 4th highest daily maximum values decreased for the vast majority of monitoring 

sites in the U.S. between 2000 and 2009. The decreasing trend is especially sharp from 2002 to 

2004, when EPA implemented the “NOX SIP Call”, a program designed to reduce summertime 

emissions of NOX in the eastern U.S., but has been relatively flat since then. 

The trends also show a modest increase in the 4th highest daily maximum values from 2009 

to 2012. This is reflected in the design value maps, which show an increase in the number of 

monitors violating the existing O3 standard in 2010-2012 relative to 2009-2011. Meteorology 

played an important role in these short-term trends. O3 concentrations tend to be higher on days 

with hot and stagnant conditions and lower on days with cool or wet conditions. According to the 

National Oceanic and Atmospheric Administration’s National Climatic Data Center (NOAANCDC), the summer of 2009 was cooler and wetter than average over most of the eastern U.S., 

while conversely the summers, of 2010, 2011, and 2012 were all much warmer than average. In 

particular, the central and eastern U.S. experienced a 2-week period of record-breaking heat in late 

June and early July of 2012, which contributed to hundreds of violations of the existing O3

standard. In contrast, the most recent climatological information available from NOAA-NCDC 

(http://www.ncdc.noaa.gov/sotc/) shows that the summer of 2013 was cooler and wetter than 

average for much of the U.S. Thus, EPA does not expect the recent increasing trend in the 4th

highest daily maximum O3 concentrations to continue in 2013. 

Figure 2-2. Trend in U.S. annual 4th highest daily maximum 8-hour O3 concentrations in 

ppb, 2000 to 2012. Solid center line represents the median value across 

monitoring sites, dashed lines represent 25th and 75th percentile values, and 

top/bottom lines represent 10th and 90th percentile values. 
2-5 

Figure 2-3. Map of 8-hour O3 design values in ppb for the 2009-2011 period. 

Figure 2-4. Map of 8-hour O3 design values in ppb for the 2010-2012 period. 
2-6 

In addition, EPA focused our analyses of welfare and ecosystem effects on a W126 O3

exposure metric in this review. The W126 metric7 is a seasonal aggregate of daytime (8:00 AM to 

8:00 PM) hourly O3 concentrations, designed to measure the cumulative effects of O3 exposure on 

plant and tree species, with units in parts per million-hours (ppm-hrs). The W126 metric uses a 

logistic weighting function to place less emphasis on exposure to low hourly O3 concentrations 

and more emphasis on exposure to high hourly O3 concentrations (Lefohn et al, 1988). 

Figure 2-5 shows the trend in annual W126 concentrations in ppm-hrs based on 933 

“trends” sites with complete data records over the 2000 to 2012 period. Figures 2-6 and 2-7 show 

maps of the 3-year average annual W126 concentrations in ppm-hrs at all U.S. monitoring sites 

for the 2009-2011 and 2010-2012 periods, respectively. The general patterns seen in these figures 

are similar to those seen in the design value metric for the existing standard. 

Figure 2-5. Trend in U.S. annual W126 concentrations in ppm-hrs, 2000 to 2012. Solid 

center line represents the median value across monitoring sites, dashed lines 

represent 25th and 75th percentile values, and top/bottom lines represent 

10th and 90th percentile values. 

 

7

 Details on the procedure used to calculate the W126 metric are provided in Chapter 4 of the welfare Risk and 

Exposure Assessment. 
2-7 

Figure 2-6. Map of 2009-2011 average annual W126 values in ppm-hrs. 

Figure 2-7. Map of 2010-2012 average annual W126 values in ppm-hrs. 
2-8 

Variations in meteorological conditions play an important role in determining ozone 

concentrations. Ozone is more readily formed on warm, sunny days when the air is stagnant. 

Conversely, ozone generation is more limited when it is cool, rainy, cloudy, or windy. EPA uses 

a statistical model to adjust for the variability in seasonal average ozone concentrations due to 

weather conditions to provide a more accurate assessment of the underlying trend in ozone 

caused by emissions (Camalier, 2007). Figure 2-8 shows the national trend in the May to 

September average of the daily maximum 8-hour ozone concentrations from 2000 to 2012 in 112 

urban locations. The dotted red line shows the trend in observed ozone concentrations at selected 

monitoring sites, while the solid blue line shows the underlying ozone trend at those sites after 

removing the effects of weather. The solid blue lines represent ozone levels anticipated under 

“typical” weather conditions and serve as a more accurate assessment of the trend in ozone due 

to changes in precursor emissions. 

Figure 2-8. Trend in the May to September mean of the daily maximum 8-hour ozone 

concentrations before (dotted red line) and after (solid blue line) adjusting for 

year-to-year variability in meteorology8. 

Figure 2-8 shows that after adjusting for the year-to-year variability in meteorology, the 

overall trend in seasonal average ozone concentrations is much smoother. The adjusted trend 

 

8

 More detailed information on these trends is available at: http://www.epa.gov/airtrends/weather.html 
2-9 

clearly shows that the NOX SIP Call program resulted in a sharp decrease in summertime ozone 

concentrations starting in 2004. The adjusted trend also indicates that ozone levels continued to 

decrease between 2004 and 2009, and while there is still some evidence of an increasing trend 

from 2009 to 2012, there is also evidence that much of the recent increase in ozone levels is due 

to meteorological conditions which were more favorable to ozone formation than normal. 

2.2 EMISSIONS AND ATMOSPHERIC CHEMISTRY 

O3 is formed by photochemical reactions of precursor gases and is not directly emitted 

from specific sources. In the stratosphere, O3 occurs naturally and provides protection against 

harmful solar ultraviolet radiation. In the troposphere, near ground level, O3 forms through 

atmospheric reactions involving two main classes of precursor pollutants: volatile organic 

compounds (VOCs) and nitrogen oxides (NOX). Carbon monoxide (CO) and methane (CH4) are 

also important for O3 formation over longer time periods (US EPA, 2013, section 3.2.2). 

Emissions of O3 precursor compounds can be divided into anthropogenic and natural 

source categories, with natural sources further divided into biogenic emissions (from vegetation, 

microbes, and animals) and abiotic emissions (from biomass burning, lightning, and geogenic 

sources). Anthropogenic sources, including mobile sources and power plants, account for the 

majority of NOX and CO emissions. Anthropogenic sources are also important for VOC 

emissions, though in some locations and at certain times of the year (e.g., southern states during 

summer) the majority of VOC emissions come from vegetation (US EPA, 2013, section 3.2.1). 

In practice, the distinction between natural and anthropogenic sources is often unclear, as human 

activities directly or indirectly affect emissions from what would have been considered natural 

sources during the preindustrial era. Thus, emissions from plants, animals, and wildfires could be 

considered either natural or anthropogenic, depending on whether emissions result from 

agricultural practices, forest management practices, lightning strikes, or other types of events 

(US EPA, 2013, sections 3.2 and 3.7.1). 

Rather than varying directly with emissions of its precursors, O3 changes in a nonlinear 

fashion with the concentrations of its precursors. NOX emissions lead to both the formation and 

destruction of O3, depending on the local quantities of NOX, VOC, radicals, and sunlight. In 

areas dominated by fresh emissions of NOX, radicals are removed, which lowers the O3

formation rate. In addition, the scavenging of O3 by reaction with NO is called “titration” and is 

often found in downtown metropolitan areas, especially near busy streets and roads, as well as in 

power plant plumes. This short-lived titration results in localized areas in which O3

concentrations are suppressed compared to surrounding areas, but which contain NO2 that 

contributes to subsequent O3 formation further downwind. Consequently, O3 response to 

reductions in NOX emissions is complex and may include O3 decreases at some times and 
2-10 

locations and increases of O3 at other times and locations. In areas with relatively low NOX

concentrations, such as those found in remote continental areas and rural and suburban areas 

downwind of urban centers, O3 production typically varies directly with NOX concentrations 

(e.g. decreases with decreasing NOX emissions). The NOx titration effect is most pronounced in 

urban core areas which have higher volume of mobile source NOx emissions from vehicles than 

do the surrounding areas. It should be noted that such locations, which are heavily NOX saturated 

(or radical limited), tend to have much lower observed O3 concentrations than downwind areas. 

As a general rule, as NOx emissions reductions occur, one can expect lower O3 values to 

increase while the higher ozone values would be expected to decrease. NOx reductions are 

expected to result in a compressed O3 distribution, relative to current conditions. 

The formation of O3 from precursor emissions is also affected by meteorological 

parameters such as the intensity of sunlight and atmospheric mixing. Major episodes of high 

ground-level O3 concentrations in the eastern United States are associated with slow-moving 

high pressure systems. High pressure systems during the warmer seasons are associated with the 

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, allowing their concentrations to build up. In addition, in 

some parts of the United States (e.g., in Los Angeles), mountain barriers limit mixing and result 

in a higher frequency and duration of days with elevated O3 concentrations. Photochemical 

activity involving precursors is enhanced during warmer seasons because of the greater 

availability of sunlight and higher temperatures (US EPA, 2013, section 3.2). 

O3 concentrations in a region are affected both by local formation and by transport of O3

and its precursors from upwind areas. O3 transport occurs on many spatial scales including local 

transport between cities, regional transport over large regions of the U.S. and international/longrange transport. In addition, O3 can be transferred into the troposphere from the stratosphere, 

which is rich in O3, through stratosphere-troposphere exchange (STE). These intrusions usually 

occur behind cold fronts, bringing stratospheric air with them and typically affect O3

concentrations in higher elevation areas (e.g. > 1500 m) more than areas at lower elevations 

(U.S. EPA, 2012, section 3.4.1.1). The role of long-range transport of ozone and other elements 

of ozone background is discussed in more detail in Section 2.4. 

2.3 AIR QUALITY CONCENTRATIONS 

Because O3 is a secondary pollutant formed in the atmosphere from precursor emissions, 

concentrations are generally more regionally homogeneous than concentrations of primary 

pollutants emitted directly from stationary and mobile sources (US EPA, 2013, section 3.6.2.1). 
2-11 

However, variation in local emissions characteristics, meteorological conditions, and topography 

can result in daily and seasonal temporal variability in ambient O3 concentrations, as well as 

local and national-scale spatial variability. 

Temporal variation in ambient O3 concentrations results largely from daily and seasonal 

patterns in sunlight, precursor emissions, atmospheric stability, wind direction, and temperature 

(US EPA, 2013, section 3.7.5). On average, ambient O3 concentrations follow well-recognized 

daily and seasonal patterns, particularly in urban areas. Specifically, daily maximum 1-hr O3

concentrations in urban areas tend to occur in mid-afternoon, with more pronounced peaks in the 

warm months of the O3 season than in the colder months (US EPA, 2013, Figures 3-54, 3-156 to 

3-157). Rural sites also follow this general pattern, though it is less pronounced in colder months 

(US EPA, 2013, Figure 3-55). With regard to day-to-day variability, median maximum daily 8-

hour average (MDA8) O3 concentrations in U.S. cities in 2007-2009 were approximately 47 ppb, 

with typical ranges between 35 to 60 ppb and the highest MDA8 concentrations above 100 ppb 

in several U.S. cities (as noted further below). 

In addition to temporal variability, there is considerable spatial variability in ambient O3

concentrations within cities and across different cities in the United States. With regard to spatial 

variability within a city, local emissions characteristics, geography, and topography can have 

important impacts. For example, as noted above, fresh NO emissions from, for example, motor 

vehicles titrate O3 present in the urban background air, resulting in an O3 gradient around 

roadways with O3 concentrations increasing as distance from the road increases (US EPA, 2013, 

section 3.6.2.1). In comparing urban areas, the ISA notes that measured O3 concentrations are 

relatively uniform and well-correlated within some cities (e.g., Atlanta) while they are more 

variable in others (e.g., Los Angeles) (US EPA, 2013, section 3.6.2.1 and Figures 3-28 to 3-36). 

With regard to variability across cities, when the ISA evaluated the distributions of 8-

hour O3 concentrations for the years 2007 to 2009 in 20 cities, the highest concentrations were 

reported in Los Angeles, with high concentrations also reported in several eastern and southern 

cities. The maximum recorded MDA8 was 137 ppb in Los Angeles, and was near or above 120 

ppb in Atlanta, Baltimore, Dallas, New York City, Philadelphia, and St. Louis (US EPA, 2013, 

Table 3-10). The pattern was similar for the 98th percentile of the distribution of MDA8 

concentrations9, with Los Angeles recording the highest 98th percentile concentration (91 ppb) 

and many eastern and southern cities reporting 98th percentile concentrations near or above 75 

ppb. In contrast, somewhat lower 98th percentile O3 concentrations were recorded in cities in the 

western United States outside of California (US EPA, 2013, Table 3-10). 

 

9

 Table 3-10 in the ISA analyzes the warm season. Therefore, the 98th percentile values would be an approximation 

of the 4th highest value. 
2-12 

Rural sites can be affected by transport of O3 or O3 precursors from upwind urban areas 

and by local anthropogenic sources such as motor vehicles, power generation, biomass 

combustion, or oil and gas operations (US EPA, 2013, section 3.6.2.2). In addition, O3 tends to 

persist longer in rural than in urban areas due to lower rates of chemical scavenging in non-urban 

environments. At higher elevations, increased O3 concentrations can also result from 

stratospheric intrusions (US EPA, 2013, sections 3.4, 3.6.2.2). As a result, O3 concentrations 

measured in some rural sites can be higher than those measured in nearby urban areas (US EPA, 

2013, section 3.6.2.2), and the ISA concludes that cumulative exposures for humans and 

vegetation in rural areas can be substantial, often higher than cumulative exposures in urban 

areas (US EPA, 2013, section 3.7.5). 

2.4 BACKGROUND O3 

One of the aspects of ozone that is unusual relative to the other pollutants with National 

Ambient Air Quality Standards (NAAQS) is that, periodically, in some locations, an appreciable 

fraction of the observed ozone results from sources or processes other than local and domestic 

regional anthropogenic emissions of ozone precursors (Fiore et al., 2002). Any ozone formed by 

processes other than the chemical conversion of local or regional ozone precursor emissions is 

generically referred to as “background” ozone. Background O3 can originate from natural 

sources of O3 and O3 precursors, as well as from manmade international emissions of O3

precursors. Natural sources of O3 precursor emissions such as wildfires, lightning, and vegetation 

can lead to O3 formation by chemical reactions with other natural sources. Another important 

component of background is O3 that is naturally formed in the stratosphere through interactions 

of ultraviolet light with molecular oxygen. Stratospheric O3 can mix down to the surface at high 

concentrations in discrete events called intrusions, especially at higher-altitude locations. The 

manmade portion of the background includes any O3 formed due to anthropogenic sources of O3

precursors emitted far away from the local area (e.g., international emissions). Finally, both 

biogenic and international anthropogenic emissions of methane, which can be chemically 

converted to O3 over relatively long time scales, can also contribute to global background O3

levels. Away from the surface, ozone can have an atmospheric lifetime on the order of weeks. As 

a result, background ozone can be transported long distances in the upper troposphere and, when 

meteorological conditions are favorable, be available to mix down to the surface and add to the 

ozone loading from non-background sources. 

As indicated in the first draft policy assessment (US EPA, 2012, sections 1.3.4 and 3), 

EPA has updated several aspects of our methodology for estimating the change in health risk and 

exposure that would result from a revision to the O3 NAAQS. First, risk estimates are now based 

on total O3 concentrations, as opposed to previous reviews which only considered risk above 
2-13 

background levels. Second, EPA is now using air quality modeling to estimate the spatial 

patterns of O3 that would result from attaining various levels of the NAAQS, as opposed to a 

quadratic rollback approach that required the estimation of a background “floor” beyond which 

the rollback would not take place. Both of these revisions have had the indirect effect of reducing 

the need for estimates of background O3 levels as part of the O3 risk and exposure assessment 

(REA). Regardless, EPA expects that a well-founded understanding of the fractional contribution 

of background sources and processes to surface O3 levels will be valuable. Accordingly, in this 

section, we briefly summarize existing results on background O3 from the ISA (US EPA, 2013, 

section 3.4) as supplemented by additional EPA modeling recently conducted for a 2007 base 

year. The summary will focus on national estimates of the: 

 

 seasonal mean background O3 values for three specific definitions of background O3, 

 relative proportion of background O3 to total O3 for the same three definitions from a 

seasonal mean perspective, 

 distributions of background O3 within a seasonal mean, 

 ratio of background O3 to total modeled ozone in the 12 REA case study areas, 

 relative proportion of background O3 concentrations to total W126 ozone, and 

 relative contribution of different components of background to total background O3. 

The definition of background O3 can vary depending upon context, but it generally refers 

to O3 that is formed by sources or processes that cannot be influenced by actions within the 

jurisdiction of concern. In the first draft policy assessment document (US EPA, 2012), EPA 

identified three specific definitions of background O3: natural background (NB), North American 

background (NAB), and United States background (USB). Natural background is the narrowest 

definition of background, and it is defined as the O3 that would exist in the absence of any 

manmade O3 precursor emissions. The other two definitions of background are based on a 

presumption that the U.S. has little influence over anthropogenic emissions outside either our 

continental or domestic borders. North American background is defined as that O3 that would 

exist in the absence of any manmade O3 precursor emissions from North America. U.S. 

background is defined as that O3 that would exist in the absence of any manmade emissions 

inside the United States. 

Each of these three definitions of background O3 requires photochemical modeling 

simulations to estimate what the residual O3 concentrations would be were the various 

anthropogenic emissions to be removed. EPA exclusively uses modeling estimates to 

characterize background, as opposed to using observed concentrations from a remote site, 

because even the most remote monitors within the U.S. can be periodically affected by U.S. 
2-14 

anthropogenic emissions. In most situations, without special monitoring it is impossible to 

determine how much of the ozone measured by a monitor originated from background sources. 

Prior to using the new 2007-based model simulations to estimate background O3 levels over the 

U.S., EPA confirmed that this modeling was able to reproduce historical O3 levels and that there 

was limited correlation between model errors and the background estimates. This evaluation is 

described more fully in the appendix (Appendix A) to this chapter. 

Previous modeling studies have estimated what background levels would be in the 

absence of certain sets of emissions by simply assessing the remaining O3 in a simulation in 

which certain emissions were removed (Zhang et al. (2011), Emery et al. (2012)). This basic 

approach is often referred to as “zero-out” modeling or “emissions perturbation” modeling. 

While the zero-out approach has traditionally been used to estimate natural background, North 

American background, and U.S. background, the methodology has an acknowledged limitation. 

It cannot answer the question of how much of the existing observed ozone results from 

background sources or processes. 

A separate modeling technique can be used to estimate the contribution of background 

ozone and other contributing source terms to total O3 within a model. This approach, referred to 

as “source apportionment” modeling, has been described and evaluated in the peer-reviewed 

literature (Dunker et al., 2002; Kemball-Cook et al., 2009). Source apportionment modeling has 

frequently been used in other regulatory settings to estimate the “contribution” to ozone of 

certain sets of emissions (EPA 2005, EPA 2011). The source apportionment technique provides a 

means of estimating the contributions of each user-identified source category to ozone formation 

in a single model simulation. This is achieved by using multiple tracer species to track the fate of 

ozone precursor emissions (VOC and NOX) and the ozone formation resulting from these 

emissions. The methodology is designed so that all ozone and precursor concentrations are 

tracked and apportioned to the selected source categories at all times without perturbing the 

inherent chemistry. The primary limitation of the source apportionment modeling is that its 

estimations of background are explicitly linked to the emissions scenarios modeled and would 

change with different emissions scenarios. 

EPA recently completed updated zero-out and source apportionment modeling for a 2007 

base year to supplement the characterization of background O3 that was provided in the ISA. 

Both of these approaches have value in assessing the potential impacts of background O3, but 

they are used separately as described in Table 2-1. 
2-15 

Estimation 

Methodology Question addressed Background Quantities Strengths and Limitations 

Zero-out 

How much ozone would remain if 

controllable emissions were 

completely removed? 

NB / NAB / 

USB 

Strength: The approach is simple 

to implement and provides an 

estimate of the lowest O3 levels 

that can be attained by 

eliminating all U.S. 

anthropogenic emissions. 

Limitation: Estimates are based 

on a counterfactual, represents a 

quantity never to occur in real 

atmosphere. Additionally, 

sensitivity approaches can be 

unreliable for evaluating mass 

contributions to O3 production 

because of non-linearity in the 

chemistry. 

Source 

Apportionment 

How much of the current ozone 

can be attributed to sources other 

than U.S. anthropogenic sources? 

Apportionmentbased USB 

Strength: Provides a direct 

estimate of the amount of O3

contributed by each source 

category while avoiding artifacts 

caused by non-linearity in the 

chemistry. 

Limitation: While this approach 

identifies important sources that 

contribute to O3, it does not 

predict quantitatively how O3 will 

respond to specific emissions 

reduction scenarios. 

Table 2-1 Comparison of the two model methodologies used to characterize background 

ozone levels. 

The key configuration elements of the updated modeling are described below; a more 

detailed description of the modeling is provided in Appendix A. The zero-out modeling was 

based on a model configuration that nested a regional-scale air quality model (CMAQ at 12 km 

horizontal grid resolution) within a global scale air quality model (GEOS-Chem at 2.0 x 2.5 
2-16 

degree horizontal grid resolution). The lateral boundary conditions from the global model were 

used as inputs for the regional simulation. Four scenarios were modeled: 

 a 2007 base case simulation which was also the basis of the air quality modeling 

performed for the 2nd draft ozone REA and is described in more detail in Appendix 4b 

of EPA (2013b) 

 a natural background run with anthropogenic ozone precursor emissions10 removed in 

both the global and regional models11 

 a North American background run with anthropogenic ozone precursor emissions 

removed across North America in both the global and regional model simulations 

 a U.S. background run with anthropogenic ozone precursor emissions removed over 

the U.S. in both the global and regional model simulations 

The source apportionment modeling was also based on a regional scale air quality model 

(CAMx at 12 km horizontal grid resolution) that used the same lateral boundary conditions from 

the 2007 base global modeling scenario. EPA used the Anthropogenic Precursor Culpability 

Assessment (APCA) tool in this analysis. The APCA tool attributes ozone production to 

manmade sources whenever ozone is determined to result from a combination of anthropogenic 

and biogenic emissions (Environ, 2011). The APCA methodology calculates natural ozone as the 

production resulting from the interaction of biogenic VOC with biogenic NOX emissions. Eleven 

separate source categories were tracked in the EPA source apportionment analysis, including five 

boundary condition terms and six in-domain sectors: 

 Boundary condition terms:12

o Northern edge 

o Eastern edge 

o Southern edge 

o Western edge 

o Top boundary 

 

10 In the global model, only emissions from natural sources were used (i.e., VOC, NOX, CO) and methane was reset 

to pre-industrial levels (700 ppb) to reflect natural contributions. In the regional modeling, the methane levels were 

left unchanged. 

11 Note that methane is not modeled as an explicit species in CMAQ but instead is treated as having a constant 

concentration. Therefore, although methane was reduced to pre-industrial levels in the global GEOS-Chem run used 

to create boundary conditions, methane was assumed to be equal to modern-day levels in the CMAQ model runs. 

Since methane reactions occur on relatively long timescales, this discrepancy is not expected to have a large impact 

on modeled background ozone levels. 

12 It should be noted that although boundary conditions are treated as part of apportionment-based USB for this 

analysis, in some cases they may be influenced by US anthropogenic emissions that are advected out of the model 

domain and recirculated back into the U.S. This is not expected to make a substantial impact on results. 
2-17 

 In-domain sectors: 

o U.S. anthropogenic emissions 

o Point sources located within the Gulf of Mexico 

o Category 3 marine vessels outside State boundaries 

o Climatologically-averaged wildfire emissions 

o Biogenic emissions 

o Canada/Mexico emissions (only those sources within the domain) 

2.4.1 Seasonal Mean Background O3 in the U.S. 

The ISA (US EPA 2013, section 3.4) previously established that background 

concentrations vary spatially and temporally and that simulated mean background concentrations 

are highest at high-elevation sites within the western U.S. Background levels typically are 

greatest over the U.S. in the spring and early summer. The updated EPA modeling focused on 

the months from April to October. Figure 2-9 displays the spatial patterns of seasonal mean 

natural background O3 as estimated by a 2007 zero-out scenario. Seasonal means are computed 

over those seven months. This figure shows the average daily maximum 8-hour O3 concentration 

(MDA8) that would exist in the absence of any anthropogenic O3 precursor emissions at monitor 

locations. As shown, seasonal mean NB levels range from approximately 15-35 ppb (i.e., +/- 1 

standard deviation) with the highest values at higher-elevation sites in the western U.S. The 

median value over these locations is 24.2 ppb, and more than 50 percent of the locations have 

natural background levels of 20-25 ppb. The highest modeled estimate of seasonal average, 

natural background, MDA8 O3 is 34.3 ppb at the high-elevation CASTNET site (Gothic) in 

Gunnison County, CO. Natural background levels are higher at these high-elevation locations 

primarily because natural stratospheric O3 impacts and international transport impacts increase 

with altitude (where O3 lifetimes are longer). 
2-18 

Figure 2-9. Map of 2007 CMAQ-estimated seasonal mean natural background O3 levels 

(ppb) from zero-out modeling. 

Figures 2-10 and 2-11 show the same information for the NAB and USB scenarios. In 

these model runs, all anthropogenic O3 precursor emissions were removed from the U.S., 

Canada, and Mexico portions of the modeling domain (NAB scenario) and then only from the 

U.S. (USB scenario). The figures show that there is not a large difference between the NAB and 

USB scenarios. Seasonal mean NAB and USB O3 levels range from 25-50 ppb, with the most 

frequent values estimated in the 30-35 ppb bin. The median seasonal mean background levels are 

31.5 and 32.7 ppb (NAB and USB, respectively). Again, the highest levels of seasonal mean 

background are predicted over the intermountain western U.S. Locations with NAB and USB 

concentrations greater than 40 ppb are confined to Colorado, Nevada, Utah, Wyoming, northern 

Arizona, eastern California, and parts of New Mexico. The 2007 EPA modeling suggests that 

seasonal mean USB concentrations are on average 1-3 ppb higher than NAB background. These 

results were similar to those reported by Wang et al. (2009). From a seasonal mean perspective, 

background levels are typically well-below the NAAQS thresholds. 
2-19 

Figure 2-10. Map of 2007 CMAQ-estimated seasonal mean North American background 

O3 levels (ppb) from zero-out modeling. 

Figure 2-11. Map of 2007 CMAQ-estimated seasonal mean United States background O3

levels (ppb) from zero-out modeling. 

2.4.2 Seasonal Mean Background O3 in the U.S. as a Proportion of Total O3

Another informative way to assess the importance of background as part of seasonal 

mean O3 levels across the U.S. is to consider the ratios of NB, NAB, and USB to total modeled 

O3 at each monitoring location. Considering the proportional impact of background allows for an 
2-20 

initial assessment of the relative importance of background and non-background sources. 

Because ozone chemistry is non-linear, one should not assume that individual perturbations (e.g., 

zero out runs) are additive in all locations. Figures 2-12 and 2-13 show the ratio of U.S. 

anthropogenic sources to total O3 using the metric of the seasonal mean MDA8 O3

concentrations as estimated by both the zero-out and source apportionment modeling 

methodologies. Recall that the terms NB, NAB, and USB are explicitly linked to the zero-out 

modeling approach. For comparison sake, in Figure 2-13 we are extending the definition of USB 

to also include the source apportionment model estimates of the O3 that is attributable to sources 

other than U.S. anthropogenic emissions. To preserve the original definition of USB, this second 

term will be hereafter referred to as “apportionment-based USB”. As noted earlier, the advantage 

of the source apportionment modeling is that all of the modeled O3 is attributed to various source 

terms without perturbing the inherent chemistry. Thus, this approach is not affected by the 

confounding occurrences of background O3 values exceeding the base O3 values as can happen 

in the zero-out modeling (i.e., background proportions > 100%). Consequently, one would 

expect the fractional background levels to be lower in the source apportionment methodology as 

a result of removing this artifact. 

When averaged over all sites, O3 from sources other than U.S. anthropogenic emissions is 

estimated to comprise 66 (zero-out) and 59 (source apportionment) percent of the total seasonal 

O3 mean. The spatial patterns of USB and apportionment-based USB are similar across the two 

modeling exercises. Background O3 is a relatively larger percentage (e.g., 70-80%) of the total 

seasonal mean O3 in locations within the intermountain western U.S. and along the U.S. border. 

In locations where O3 levels are generally higher, like California and the eastern U.S., the 

seasonal mean background fractions are relatively smaller (e.g., 40-60%). The additional 2007 

modeling confirms that background ozone, while generally not approaching levels of the ozone 

standard, can comprise a considerable fraction of total seasonal mean ozone across the U.S. 

2.4.3 Daily Distributions of Background O3 within the Seasonal Mean 

 As a first-order understanding, it is valuable to be able to characterize seasonal mean 

levels of background O3. However, it is well established that background levels can vary 

substantially from day-to-day within the seasonal mean. From an implementation perspective, 

the values of background O3 on possible exceedance days are a more meaningful consideration. 

The first draft policy assessment (US EPA, 2012) considered this issue in detail, via summaries 

of the existing 2006 zero-out modeling (Henderson et al., 2012), and concluded that “results 

suggest that background concentrations on the days with the highest total O3 concentrations are 

not dramatically higher than typical seasonal average background concentrations.” Based on this 

finding, EPA determined that “anthropogenic sources within the U.S. are largely responsible for 
2-21 

4th highest 8-hour daily maximum O3 concentrations.” The recent EPA modeling using a 2007 

base year and the two distinct modeling methodologies described above, supports the finding 

from the previous 2006-based modeling analyses. That is, the highest modeled O3 site-days tend 

to have background O3 levels similar to mid-range O3 days. Figure 2-14 and 2-15 show the 

distribution of daily MDA8 apportionment-based USB levels (absolute magnitudes and relative 

fractions, respectively) from the source apportionment simulation13. Again, the 2007 modeling 

shows that the days with highest O3 levels have similar distributions (i.e., means, inter-quartile 

ranges) of background levels as days with lower values, down to approximately 40 ppb. As a 

result, the proportion of total O3 that has background origins is smaller on high O3 days (e.g., 

days > 60 ppb) than on the more common lower O3 days that tend to drive seasonal means. This 

helps put the results from section 2.4.2 into better context. For example, for site-days in which 

base O3 is between 70-75 ppb, the source apportionment modeling estimates that approximately 

37 percent of those O3 levels originate from sources other than U.S. anthropogenic emissions 

(i.e., apportionment-based USB). Figure 2-15 also indicates that there are cases in which the 

model predicts much larger background proportions, as shown by the upper outliers in the figure. 

These infrequent episodes usually occur in relation to a specific event, and occur more often in 

specific geographical locations, such as at high elevations or wildfire prone areas during the local 

dry season. 

It should be noted here that EPA has policies for treatment of air quality monitoring data 

affected by these types of events. EPA’s exceptional events policy allows exclusion of certain air 

quality monitoring data from regulatory determinations if a State adequately demonstrates that an 

exceptional event has caused the exceedance or violation of a NAAQS. In addition, Section 

179B of the CAA also provides for treatment of air quality data from international transport 

when an exceedance or violation of a NAAQS would not have occurred but for the emissions 

emanating from outside of the United States. 

 

13 Similar plots from the zero-out modeling for natural background, North American background, and U.S. 

background are provided in Appendix A. 
2-22 

Figure 2-12. Map of site-specific ratios of U.S. background to total seasonal mean O3 based 

on 2007 CMAQ zero-out modeling. 

Figure 2-13. Map of site-specific ratios of apportionment-based U.S. background to 

seasonal mean O3 based on 2007 CAMx source apportionment modeling. 
2-23 

Figure 2-14. Distributions of absolute estimates of apportionment-based U.S. Background 

(all site-days), binned by modeled MDA8 from the 2007 source 

apportionment simulation. 

Figure 2-15. Distributions of the relative proportion of apportionment-based U.S. 

Background to total O3 (all site-days), binned by modeled MDA8 from the 

2007 source apportionment simulation. 
2-24 

2.4.4 Proportion of Background O3 in 12 Urban Case Study Areas 

 This section presents estimates of the overall fraction of O3 that is estimated to result 

from background sources or processes in each of the 12 urban case study areas considered in the 

epidemiological-based risk assessment of the REA (US EPA 2014, Chapter 7). The results are 

based on the recent EPA 2007 source apportionment modeling. Table 2-1 summarizes the 

estimated ratios of sources other than U.S. anthropogenic emissions (i.e., apportionment-based 

USB) to total seasonal mean MDA8 O3 in each of the 12 urban case study areas. The table shows 

that the fractional contributions from sources other than anthropogenic emissions within the U.S. 

can range from 43 to 66 percent across these 12 urban areas. These fractions are consistent with 

the national ratios summarized in section 2.4.2, although the fractions of background are 

generally smaller at urban sites than at rural sites. 

 As shown in section 2.4.3, the background-to-total ratios are smaller on days with high 

modeled O3 (i.e., days that may exceed the level of the NAAQS). Table 2-2 provides the 

fractional contributions from apportionment-based USB, only considering days in which base 

model MDA8 O3 was greater than 60 ppb. As expected, the fractional background contributions 

are smaller, ranging from 31 to 55 percent. 

 Rather than taking the fractions of the seasonal means (as in Table 2-1), Table 2-3 

displays the mean and median daily MDA8 background fractions. These metrics may be more 

appropriate for application to health studies. The fractional contributions to background 

calculated via this approach are very similar to the Table 2-1 calculations. For completeness, we 

also provide USB ratios based on the zero-out modeling for the 12 cities (see Table 2-4). The 

results are similar to the source apportionment findings (Table 2-1), though the zero-out 

technique provides slightly higher background proportions, as expected. It should be noted that 

all fractional contributions are based on recent conditions from 2007. These ratios would be 

expected to change as anthropogenic emissions and O3 levels are lowered. 

Based on the source apportionment modeling for these 12 areas, there is evidence that 

background levels comprise a non-negligible fraction of the total ozone observed within these 

locations. However, for site-days in which model MDA8 ozone exceeds 60 ppb, ozone formed 

from U.S. anthropogenic emissions comprise a larger fraction of the total ozone in 11 of the 12 

areas (all but Denver). The major metropolitan areas in the eastern U.S. (e.g., Atlanta, New York 

City, Philadelphia) are less influenced by background sources than a higher-elevation, western 

U.S., location like Denver. Even in Denver, though, U.S. anthropogenic emissions have a large 

influence on total ozone (45 percent). 
2-25 

Table 2-2 Seasonal mean MDA8 O3 (ppb), seasonal mean apportionment-based USB 

contribution (ppb), and fractional apportionment-based USB contribution to 

total O3 (all site-days) in the 12 REA urban case study areas (%). 

Table 2-3 Seasonal mean MDA8 O3 (ppb), seasonal mean apportionment-based USB 

contribution (ppb), and fractional apportionment-based USB contribution to 

total O3 (site-days > 60 ppb) in the 12 REA urban study areas (%). 

Table 2-4 Fractional contribution of apportionment-based USB in the 12 REA urban 

study areas (%), using the means and medians of daily MDA8 fractions 

(instead of fractions of seasonal means). 

Table 2-5 Seasonal mean MDA8 O3 (ppb), seasonal mean USB (ppb), and USB/Total ratio 

(all site-days) in the 12 REA urban case study areas (%). 

2.4.5 Influence of Background O3 on W126 levels 

 EPA also conducted a limited assessment of the impacts of background O3 sources on the 

W126 metric. The W126 metric (LeFohn et al., 1988) is a cumulative peak-weighted index 

designed to estimate longer-term effects of daytime ozone levels on sensitive vegetation and 

ecosystems. EPA used the 2007 zero-out modeling to assess NB, NAB, and USB influences at 

All days, CAMx ATL BAL BOS CLE DEN DET HOU LA NYC PHI SAC STL

Model MDA8 seasonal mean 59.3 54.4 43.0 48.9 47.3 39.1 48.5 51.1 45.4 48.7 46.4 49.8

Model MDA8 seasonal mean

from emissions other than

U.S. anthropogenic sources

25.3 25.9 26.2 25.7 31.3 23.3 27.0 29.1 24.5 24.2 29.7 24.3

Fractional contribution from

background 0.43 0.48 0.61 0.53 0.66 0.60 0.56 0.57 0.54 0.50 0.64 0.49

Only days w/ base

MDA8 > 60 ppb ATL BAL BOS CLE DEN DET HOU LA NYC PHI SAC STL

Model MDA8 seasonal mean 74.0 75.3 70.7 72.0 67.5 68.9 70.3 74.4 74.1 74.0 68.3 70.0

Model MDA8 seasonal mean

from emissions other than

U.S. anthropogenic sources

25.4 23.7 24.4 25.4 37.3 24.4 28.0 31.9 23.5 22.9 32.1 25.4

Fractional contribution from

background 0.34 0.31 0.35 0.35 0.55 0.35 0.40 0.43 0.32 0.31 0.47 0.36

ATL BAL BOS CLE DEN DET HOU LA NYC PHI SAC STL

Mean of daily MDA8

background fractions 0.46 0.53 0.68 0.58 0.69 0.64 0.59 0.61 0.61 0.56 0.67 0.52

Median of daily MDA8

background fractions 0.43 0.51 0.73 0.54 0.69 0.66 0.59 0.60 0.63 0.54 0.66 0.49

All days, CMAQ ATL BAL BOS CLE DEN DET HOU LA NYC PHI SAC STL

Model MDA8 seasonal mean 58.6 55.6 45.2 51.8 57.1 43.5 49.4 54.8 47.7 50.5 51.9 52.6

Model MDA8 seasonal mean

USB

30.0 29.9 28.5 31.6 42.2 31.7 33.0 33.3 29.1 29.4 34.4 32.0

Ratio of USB/Total MDA8 O3 0.51 0.54 0.63 0.61 0.74 0.73 0.67 0.61 0.61 0.58 0.66 0.61
2-26 

four sample locations: Atlanta GA, Denver CO, Farmington NM, and Riverside CA. Each of the 

four analyses locations had relatively high observed values of W126 in 2007, as averaged over 

all sites in the area: Atlanta (25.1 ppm-hrs), Denver (19.6 ppm-hrs), Farmington (20.2 ppm-hrs), 

and Riverside (36.0 ppm-hrs). 

 As discussed above, in the current review, EPA is supplementing the counterfactual 

assessment used in previous reviews (zero out modeling) with analyses that estimate the portion 

of the existing ozone that is due to background sources (source apportionment). This has 

important ramifications for assessing the influence of background on W126 concentrations, 

because of the non-linear weighting function used in the metric, which emphasizes high ozone 

hours (e.g., periods in which ozone is greater than ~60 ppb). As an example, consider a sample 

site in the intermountain western U.S. region with very high modeled estimates of U.S 

background (e.g., seasonal mean of 45 ppb with some days as high as 65 ppb). Even at this high 

background location, the USB simulations estimate annual W126 (USB) values that are quite 

low, on the order of 3 ppm-hrs. Sites in the domain with lower U.S. background levels have even 

smaller USB W126 values, on the order of the 1 ppm-hrs, which is consistent with values 

mentioned in past reviews (US EPA, 2007). Using the counterfactual scenarios, background 

ozone has a relatively small impact on W126 levels across the U.S.

 However, because of the non-linear weighting function used in the W126 calculation, the 

sum of the W126 from the USB scenario and the W126 resulting from U.S. anthropogenic 

sources will not equal the total W126. In most cases, the sum of those two components will be 

substantially less than total W126. As a result, EPA believes it is more informative to estimate 

the fractional influence of background ozone to W126 levels. Using a methodology that is 

described in more detail in Appendix A, EPA considered the fractional influence of background 

ozone on annual W126 levels in four locations. The fractional influence methodology utilizes the 

2007 zero-out modeling but places higher weights on background fractions on days that are 

going to contribute most substantially to the yearly W126 value. Figure 2-16 shows the results. 

Based on the fractional influence methodology, natural background sources are estimated to 

contribute 29-50% of the total modeled W126 with the highest relative influence in the 

intermountain western U.S. (i.e., Farmington NM) and the lowest relative influence in the eastern 

U.S. (i.e., Atlanta). U.S. background is estimated to contribute 37-65% of the total modeled 

W126. The proportional impacts of background are slightly less for the W126 metric than for 

seasonal mean MDA8 (discussed in section 2.4.2), because of the sigmoidal weighting function 

that places more emphasis on higher ozone days when background fractions are generally lower.

 The key conclusion from this cursory analysis is that background ozone can comprise a 

non-negligible portion of current W126 levels across the U.S. These fractional influences are 

greatest in the intermountain western U.S. and are slightly smaller than the seasonal mean 
2-27 

MDA8 metric. This conclusion was also reached in a separate analysis recently completed by 

Lapina et al. (2014). 

Figure 2-16. Fractional influence of background sources to W126 levels in four sample 

locations. Model estimates based on 2007 CMAQ zero-out modeling. 

2.4.6 Estimated Magnitude of Individual Components of Background O3

 To provide a fuller characterization of background O3 levels, it is useful to develop an 

understanding of the relative contributions of various background elements to total background 

O3. This section will utilize the supplemental 2007 air quality modeling estimates to consider the 

relative contribution of specific elements of background O3. Several background elements were 

isolated in either the zero-out or source apportionment modeling. Appendix A provides more 

detail on these analyses. In conjunction with the previous analyses summarized in the ISA, some 

broad characterizations of the individual components of background O3 can be developed. 

 The recent 2007 EPA modeling confirms the importance of methane emissions and 

international ozone precursor emissions in contributing to background O3. Methane has an 

atmospheric lifetime of about a decade and can be an important contributor to ozone on longer 

time scales. Anthropogenic methane emission sources include agriculture, coal mines, landfills, 

and natural gas and oil systems. The difference between the NAB and NB zero-out scenarios 

provides an estimate of contributions from international anthropogenic emissions and 

anthropogenic methane, which is modeled by reducing model concentrations from present-day 

values to pre-industrial levels. The ISA (US EPA, 2013, section 3.4) estimated that roughly half 

of the difference between the NB and NAB scenarios resulted from the removal of anthropogenic 

0%

10%

20%

30%

40%

50%

60%

70%

80%

90%

100%

Farmington Denver Riverside Atlanta

Percent contribution from background to W126

NB

NAB

USB
2-28 

methane emissions and that the other half resulted from international anthropogenic emissions of 

shorter-lived O3 precursors (e.g., NOX and VOC). Figure 2-17 shows the difference in seasonal 

mean MDA8 O3 levels between the NB and NAB scenarios. North American seasonal mean 

background is 6-15 ppb higher than comparable natural background levels. The most frequent 

increment is an 8-10 ppb increase when the methane is increased and the non-North American 

emissions are added. These estimates of seasonal-mean external enhancement are similar to 

previous estimates summarized in the ISA (e.g., Fiore et al., 2009; Zhang et al., 2011). It is not 

possible via the EPA 2007 modeling to parse out what fraction of this change is due to emissions 

outside of North America, as opposed to what fraction is due to anthropogenic methane 

emissions, but the modeling suggests that any control measures to reduce emissions from either 

of these terms have the potential to contribute in an important way to reduce average background 

O3 levels in the U.S. 

 The difference between the NAB and USB scenarios is easier to interpret as it only 

involves one change, the inclusion of anthropogenic emissions from the in-domain portion of 

Canada and Mexico. These emissions (not shown here, but depicted in Appendix A) contribute 

less than 2 ppb to the seasonal mean MDA8 O3 levels over most of the U.S. There are 70 sites, 

near an international border, where the modeling estimates Canadian/Mexican seasonal average 

impacts of 2-4 ppb. Peak single-day MDA8 impacts from these specific international emissions 

sources can approach 25 ppb (e.g., San Diego, Buffalo NY). 

 

Figure 2-17. Differences in seasonal mean O3 (ppb) between the NAB and NB scenarios. 
2-29 

 Eleven separate source categories were tracked in the source apportionment modeling, 

including five boundary condition terms (East, South, West, North, and top) and six emissions 

sectors within the domain. The contributions of each of these terms is provided in the Appendix 

and summarized below. At most locations, the five model boundary terms contributed an 

aggregate 40-60 percent of the total seasonal mean MDA8 O3 across the U.S. The highest 

proportional impacts from the boundary conditions are along the coastlines and the 

intermountain West. The O3 entering the model domain via the boundary conditions can have a 

variety of origins including: a) natural sources of O3 and precursors emanating from outside the 

domain, b) anthropogenic sources of O3 precursors emanating from outside the domain, and c) 

some fraction of U.S. emissions (natural and anthropogenic) which exit the regional model 

domain but get re-imported into the domain via synoptic-scale recirculation. Accordingly, it is 

not possible to relate the boundary condition contribution to any specific background element. 

The single largest sector that was tracked in the source apportionment modeling was U.S. 

anthropogenic emissions. Figure 2-18 shows the contributions from this sector to seasonal mean 

MDA8 O3 levels. Over most of the U.S. this term contributes 40-60 percent to the total seasonal 

mean O3. As discussed in section 2.4.3, these contributions are even higher when only high O3

days are considered. International shipping emissions, as well as fires and other biogenic 

emissions also contribute in a non-negligible way to background O3 over the U.S. The key 

finding from this analysis is that air quality planning efforts to reduce anthropogenic methane 

emissions and international NOX/VOC emissions (e.g., migrating from Asia, Canada, and 

Mexico; and from commercial shipping) have the potential to lower background O3 levels. 

 
2-30 

Figure 2-18. Percent contribution of U.S. anthropogenic emissions to total seasonal mean 

MDA8 O3 levels, based on 2007 source apportionment modeling. 

2.4.7 Summary 

For a variety of reasons, it is challenging to present a comprehensive summary of all the 

components and implications of background O3. In many forums the term “background” is used 

generically and the lack of specificity can lead to confusion as to what sources are being 

considered. Additionally, it is well established that the impacts of background sources can vary 

greatly over space and time which makes it difficult to present a simple summary of background 

O3 levels. Further, background O3 can be generated by a variety of processes, each of which can 

lead to differential patterns in space and time, and which often have different regulatory 

ramifications. Finally, background O3 is difficult to measure and thus, typically requires air 

quality modeling which has inherent uncertainties and potential errors and biases. 

That said, EPA believes the following concise and three step summary of the implications 

of background O3 on the NAAQS review is appropriate, as based on previous modeling exercises 

and the more recent EPA analyses summarized herein. First, background O3 exists and can 

comprise a considerable fraction of total seasonal mean MDA8 O3 and W126 across the U.S. Air 

quality models can estimate the fractional contribution of background sources to total O3 in an 

individual area. The largest absolute values of background (NB, NAB, USB, or apportionment-
2-31 

based USB) are modeled to occur at locations in the intermountain western U.S. and are 

maximized in the spring and early summer seasons. Second, the modeling indicates that U.S. 

anthropogenic emission sources are the dominant contributor to the majority of modeled O3

exceedances of the NAAQS. Higher O3 days generally have smaller fractional contributions 

from background. This finding indicates that the relative importance of background O3 would 

increase were O3 concentrations to decrease with a lower level of the O3 NAAQS. Third and 

finally, while the majority of modeled O3 exceedances have local and domestic regional 

emissions as their primary cause, there can be events where O3 levels approach or exceed 60-75 

ppb due to the influence of background sources. These events are relatively infrequent and EPA 

has policies that could allow for the exclusion of air quality monitoring data affected by these 

types of events from design value calculations. 
2-32 

2.5 REFERENCES 

Camalier, L.; Cox, W.; Dolwick, P. (2007). The Effects of Meteorology on Ozone in Urban Areas and their use in 

Assessing Ozone Trends. Atmos Environ 41: 7127-7137. 

Dunker, A.M.; Yarwood, G; Ortmann, J.P.; Wilson, G.M. (2002). Comparison of source apportionment and source 

sensitivity of ozone in a three-dimensional air quality model. Environmental Science & Technology 36: 

2953-2964. 

Emery, C; Jung, J; Downey, N; Johnson, J; Jimenez, M; Yarwood, G; Morris, R. (2012). Regional and global 

modeling estimates of policy relevant background ozone over the United States. Atmos Environ 47: 206-

217. http://dx.doi.org/10.1016/j.atmosenv.2011.11.012. 

Lapina K., Henze D.K., Milford J.B., Huang M., Lin M., Fiore A.M., Carmichael G., Pfister G.G., Bowman K. 

(2014). Assessment of source contributions to seasonal vegetative exposure to ozone in the U.S. Journal of 

Geophysical Research-Atmospheres, 119: DOI: 10.1002/2013JD020905. 

Lefohn, A. S.; Laurence, J. A.; Kohut, R. J. (1988). A comparison of indices that describe the relationship between 

exposure to ozone and reduction in the yield of agricultural crops. Atmos. Environ. 22: 1229-1240. 

Henderson, B.H., Possiel, N., Akhtar, F., Simon, H.A. (2012). Regional and Seasonal Analysis of North American 

Background Ozone Estimates from Two Studies. Available on the Internet at: 

http://www.epa.gov/ttn/naaqs/standards/ozone/s_o3_td.html 

Kemball-Cook, S.; Parrish, D.; Ryerson, T; Nopmongcol, U.; Johnson, J.; Tai, E.; Yarwood, G. (2009). 

Contributions of regional transport and local sources to ozone exceedances in Houston and Dallas: 

Comparison of results from a photochemical grid model to aircraft and surface measurements. Journal of 

Geophysical Research-Atmospheres, 114: D00F02. DOI: 10.1029/2008JD010248. 

U.S. Environmental Protection Agency (2005). Technical Support Document for the Final Clean Air Interstate Rule 

Air Quality Modeling. Office of Air Quality Planning and Standards, Research Triangle Park, NC, 285pp. 

http://www.epa.gov/cair/technical.html. 

U.S. Environmental Protection Agency. (2007). Review of the National Ambient Air Quality Standards for Ozone: 

Policy Assessment of Scientific and Technical Information - OAQPS Staff Paper, U.S. Environmental 

Protection Agency, Research Triangle Park, NC. EPA 452/R-07-007. 

U.S. Environmental Protection Agency (2011). Air Quality Modeling Final Rule Technical Support Document. 

Office of Air Quality Planning and Standards, Research Triangle Park, NC, 363pp. 

http://www.epa.gov/airtransport/CSAPR/techinfo.html. 

U.S. Environmental Protection Agency. (2012). Policy Assessment for the Review of the Ozone National Ambient 

Air Quality Standards, First External Review Draft, U.S. Environmental Protection Agency, Research 

Triangle Park, NC. EPA 452/P-12-002. 

U.S. Environmental Protection Agency. (2013). Integrated Science Assessment for Ozone and Related 

Photochemical Oxidants, U.S. Environmental Protection Agency, Research Triangle Park, NC. 

EPA/600/R-10/076. 

U.S. Environmental Protection Agency. (2014). Health Risk and Exposure Assessment for Ozone, Second External 

Review Draft, U.S. Environmental Protection Agency, Research Triangle Park, NC. EPA-452/P-14-004a. 

Wang, HQ; Jacob, DJ; Le Sager, P; Streets, DG; Park, RJ; Gilliland, AB; van Donkelaar, A. (2009). Surface ozone 

background in the United States: Canadian and Mexican pollution influences. Atmos Environ 43: 1310-

1319. http://dx.doi.org/10.1016/j.atmosenv.2008.11.036. 
2-33 

Zhang, L; Jacob, DJ; Downey, NV; Wood, DA; Blewitt, D; Carouge, CC; Van donkelaar, A; Jones, DBA; Murray, 

LT; Wang, Y. (2011). Improved estimate of the policy-relevant background ozone in the United States 

using the GEOS-Chem global model with 1/2° × 2/3° horizontal resolution over North America. Atmos 

Environ 45: 6769-6776. http://dx.doi.org/10.1016/j.atmosenv.2011.07.054. 
3-1 

3 Adequacy of the Current Primary Standard 

This chapter presents staff’s considerations and conclusions regarding the adequacy of 

the current primary O3 NAAQS. In doing so, we pose the following overarching question: 

As discussed more fully in this chapter, staff reaches the conclusion that the available 

evidence and exposure and risk information clearly calls into question the adequacy of public 

health protection provided by the current primary standard. This evidence and information 

provides strong support for the occurrence of a range of adverse respiratory effects, and 

mortality, under air quality conditions that would meet the current standard. Based on the 

analyses in the HREA, we conclude that the exposures and risks projected to remain upon 

meeting the current standard are indicative of risks that can reasonably be judged to be important 

from a public health perspective. Thus, staff concludes that the evidence and information 

provides strong support for giving consideration to revising the current primary standard in order 

to provide increased public health protection against an array of adverse health effects that range 

from decreased lung function and respiratory symptoms to more serious indicators of morbidity 

(e.g., including emergency department visits and hospital admissions), and mortality. The 

remainder of this chapter discusses the evidence and exposure/risk information, and the 

considerations and conclusions based on that evidence and information, supporting staff’s 

overarching conclusion regarding the adequacy of the current primary O3 standard. 

In addressing the overarching question for this chapter, we pose a series of more specific 

questions, as discussed in sections 3.1 through 3.4 below. Section 3.1 presents our consideration 

of the available scientific evidence (i.e., evidence-based considerations) about the health effects 

associated with short- and long-term O3 exposures. Section 3.2 presents our consideration of 

available estimates of O3 exposures and health risks (exposure- and risk-based considerations). 

Section 3.3 discusses the advice and recommendations that we have received from the CASAC 

on the first draft O3 PA, and on documents from previous reviews of the O3 NAAQS. Section 3.4 

revisits the overarching question of this section, and presents staff’s conclusions regarding the 

adequacy of the current primary O3 NAAQS. 

Does the currently available scientific evidence and exposure/risk information, as 

reflected in the ISA and HREA, support or call into question the adequacy of the 

current O3 standard?
3-2 

3.1 EVIDENCE-BASED CONSIDERATIONS 

This section presents our consideration of the available scientific evidence with regard to 

the adequacy of the current O3 standard. Our approach, as summarized in section 1.3.1 above, is 

based on the full body of evidence in this review. We use information from the full evidence 

base to characterize our confidence in the extent to which O3-attributable effects occur, and the 

extent to which such effects are adverse, over the ranges of O3 exposure concentrations evaluated 

in controlled human exposure studies and over the distributions of ambient O3 concentrations in 

locations where epidemiologic studies have been conducted. In doing so, we recognize that the 

available health effects evidence reflects a continuum from relatively high O3 concentrations, at 

which scientists generally agree that adverse health effects are likely to occur, through lower 

concentrations, at which the likelihood and magnitude of a response become increasingly 

uncertain. 

Section 3.1.1 summarizes a mode of action framework for understanding the effects of 

both short- and long-term O3 exposures, based on Chapter 5 of the ISA (U.S. EPA, 2013). 

Section 3.1.2 presents our consideration of the evidence for health effects attributable to shortterm and long-term O3 exposures. Section 3.1.3 discusses the adversity of the effects. Section 

3.1.4 presents our consideration of evidence with regard to concentrations associated with health 

effects and section 3.1.5 presents our consideration of the public health implications of exposures 

to O3, including the adversity of effects and evidence for at-risk populations and lifestages.1

 

3.1.1 Modes of Action 

Our consideration of the evidence of effects attributable to short-and long-term exposures 

and the factors that increase risk in populations and lifestages builds upon evidence about the 

modes of action by which O3 exerts effects (U.S. EPA, 2013; section 5.3). Mode of action refers 

to a sequence of key events and processes that result in a given toxic effect; elucidation of 

mechanisms provides a more detailed understanding of these key events and processes. The 

purpose of this section is to describe the key events and pathways that contribute to health effects 

resulting from both short-term and long-term exposures to O3. The extensive research carried out 

 

1

 Here, as in the ISA, the term “at-risk population” is used to encompass populations or lifestages that have a greater 

likelihood of experiencing health effects related to exposure to an air pollutant due to a variety of factors; other 

terms used in the literature include susceptible, vulnerable, and sensitive. These factors may be intrinsic, such as 

genetic factors, lifestage, or the presence of preexisting diseases, or they may be extrinsic, such as socioeconomic 

status (SES), activity pattern and exercise level, or increased pollutant exposures (U.S. EPA, 2013, p. lxx, 8-1, 8-2). 

The courts and the Act’s legislative history refer to these at-risk subpopulations as “susceptible” or “sensitive” 

populations. See, e.g., American Lung Ass’n v. EPA, 134 F. 3d 388, 389 (D.C. Cir. 1998) (“NAAQS must protect 

not only average health individuals, but also ‘sensitive citizens’ – children, for example, or people with asthma, 

emphysema, or other conditions rendering them particularly vulnerable to air pollution” (quoting S. Rep. No. 91-

1196 at 10). 
3-3 

over several decades in humans and animals has yielded numerous studies on mechanisms by 

which O3 exerts its effects. It is well-understood that secondary oxidation products, which form 

as a result of O3 exposure, initiate numerous responses at the cellular, tissue and whole organ 

level of the respiratory system. These responses include the activation of neural reflexes, 

initiation of inflammation, alteration of barrier epithelial function, sensitization of bronchial 

smooth muscle, modification of lung host defenses, and airways remodeling, as discussed below. 

These key events have the potential to affect other organ systems such as the cardiovascular 

system. It has been proposed that secondary oxidation products, which are bioactive and 

cytotoxic in the respiratory system, are also responsible for systemic effects. Recent studies in 

animal models show that inhalation of O3 results in systematic oxidative stress. 

Figure 3.1 below, adapted from Figure 5-8 of the ISA (ISA, Section 5.3.10, U.S. EPA, 

2013), shows key events in the toxicity pathway of O3 that are described in more detail below. 

The initial key event in the toxicity pathway of O3 is the formation of secondary oxidation 

products in the respiratory tract (ISA, section 5.3, U.S. EPA, 2013). This mainly involves direct 

reactions with components of the extracellular lining fluid (ELF). Although the ELF has inherent 

capacity to quench (based on individual antioxidant capacity), this capacity can be overwhelmed, 

especially with exposure to elevated concentrations of O3.2 The resulting secondary oxidation 

products transmit signals to the epithelium, pain receptive nerve fibers and, if present, immune 

cells (i.e., eosinophils, dendritic cells and mast cells) involved in allergic responses. Thus, the 

effects of O3 are mediated by components of ELF and by the multiple cell types found in the 

respiratory tract. Further, oxidative stress3 is an implicit part of this initial key event. 

 

2

 The ELF is a complex mixture of lipids (fats), proteins, and antioxidants that serve as the first barrier and target for 

inhaled O3. The quenching ability of antioxidant substances present in the ELF appear in most cases to limit 

interaction of O3 with underlying tissues and to prevent penetration of O3 deeper into the lung. However, as the ELF 

thickness decreases and becomes ultra thin in the alveolar region, it may be possible for direct interaction of O3 with 

the underlying epithelial cells to occur. The formation of secondary oxidation products is likely related to the 

concentration of antioxidants present and the quenching ability of the lining fluid. 

3

 Oxidative stress reflects an imbalance between the systemic manifestation of reactive oxygen species, such as 

superoxide, and a biological system's ability to readily detoxify the reactive intermediates or to repair the resulting 

damage. 
3-4 

Figure 3-1. Modes of action/possible pathways underlying the health effects resulting from 

inhalation exposure to O3. (Adapted from U.S. EPA, 2013, Figure 5-8)

Another key event in the toxicity pathway of O3 is the activation of neural reflexes which 

lead to lung function decrements. Evidence is accumulating that secondary oxidation products 

are responsible for this effect. Different receptors on bronchial sensory nerves (i.e., C-fibers) 

have been shown to mediate separate effects of O3 on pulmonary function. For example, pain 

(i.e., nociceptive) sensory nerves are involved in the involuntary truncation of inspiration which 

results in decreases in FVC, FEV1, tidal volume and pain upon deep inspiration. Ozone exposure 

also results in activation of vagal sensory nerves and a mild increase in airway obstruction 

measured as increased sRaw. Activation of neural reflexes also results in extrapulmonary effects 

such as slow resting heart rate (i.e., bradycardia). 

Initiation of inflammation is also a key event in the toxicity pathway of O3. Secondary 

oxidation products, as well as cell signaling molecules (i.e., chemokines and cytokines) from 

airway epithelial cells and white blood cells (i.e., macrophages), have been implicated in the 

initiation of inflammation. Airways neutrophilia has been demonstrated in bronchoalveolar 

lavage fluid (BALF), mucosal biopsy and induced sputum samples. Influx of other cells (i.e., 

mast cells, monocytes and macrophages) also occur. Inflammation further contributes to 

Mode of Action/Possible Pathways

Formation of secondary oxidation products

Ozone + Respiratory Tract

Systemic inflammation and

oxidative/nitrosative stress

Activation

of neural       

reflexes

Airways

remodeling

Sensitization

of bronchial

smooth muscle

Alteration

of epithelial

barrier

function

Initiation of

inflammation    

Modification

of innate and

adaptive

immunity

Extrapulmonary Effects

Decrements in pulmonary function

Pulmonary inflammation/oxidative stress

Increases in airways permeability

Airways hyperresponsiveness

Exacerbation/induction of asthma

Decreased host defenses

Epithelial metaplasia and fibrotic changes

Altered lung development
3-5 

O3-mediated oxidative stress. It should be noted that inflammation, as measured by airways 

neutrophilia, is not correlated with decrements in pulmonary function as measured by 

spirometry. 

A fourth key event in the toxicity pathway of O3 is alteration of epithelial barrier 

function. Increased permeability4

 occurs as a result of damage to tight junctions between 

epithelial cells subsequent to O3-induced injury and inflammation. It may play a role in allergic 

sensitization and in airway hyperresponsiveness (AHR). Genetic susceptibility has been 

associated with this pathway. 

A fifth key event in the toxicity pathway of O3 is the sensitization of bronchial smooth 

muscle. Airway hyperresponsiveness, or increased bronchial reactivity, can be both a rapidly 

occurring and a persistent response. The mechanisms responsible for AHR are not wellunderstood. Tachykinins, peptides that can excite neurons and cause smooth muscle contraction, 

and the secondary oxidation products of O3 have been proposed as mediators of the early 

response and inflammation-derived products have been proposed as mediators of the later 

response. Other chemical signaling molecules (i.e., cytokines and chemokines) have been 

implicated in the AHR response to O3 in animal models. Antioxidants may confer protection. 

A sixth key event in the toxicity pathway of O3 is the modification of innate/adaptive 

immunity. While the majority of evidence for this key event comes from animal studies, there 

are several studies suggesting that this pathway may also be relevant in humans. Ozone exposure 

of human subjects resulted in recruitment of activated innate immune cells to the airways. 

Animal studies further linked O3-mediated activation of the innate immune system to the 

development of nonspecific AHR, demonstrated an interaction between allergen and O3 in the 

induction of nonspecific AHR, and found that O3 acted as an adjuvant for allergic sensitization 

through the activation of both innate and adaptive immunity. These studies provide evidence that 

O3 can alter host immunologic response and lead to immune system dysfunction. These 

mechanisms may underlie the exacerbation and induction of asthma, as well as decreases in lung 

host defense. 

Another key event in the toxicity pathway of O3 is airways remodeling. Persistent 

inflammation and injury, which are observed in animal models of chronic and intermittent 

 

4

 Cells in epithelium are very densely packed together, leaving very little intercellular space. All epithelial cells rest 

on a basement membrane, a thin sheet of fibers that acts as scaffolding on which epithelium can grow and regenerate 

after injuries. Epithelial tissue is innervated but avascular; it must be nourished by substances diffusing from the 

blood vessels in the underlying tissue. Injury to epithelial cells, such as caused by oxidative stress, can cause the 

epithelium to become more permeable to substances in the underlying vasculature. 
3-6 

exposure to O3, are associated with morphologic changes such as mucous cell metaplasia5 of 

nasal epithelium, bronchiolar metaplasia of alveolar ducts and fibrotic changes in small airways 

(see Section 7.2.3 of the ISA, U.S. EPA, 2013). Mechanisms responsible for these responses are 

not well-understood. However, a recent study in mice demonstrated a key role for a signaling 

pathway in the deposition of collagen in the airway wall following chronic intermittent exposure 

to O3. Chronic intermittent exposure to O3 has also been shown to result in effects on the 

developing lung and immune system. 

Systemic inflammation and vascular oxidative/nitrosative stress are also key events in the 

toxicity pathway of O3. Extrapulmonary effects of O3 occur in numerous organ systems, 

including the cardiovascular, central nervous, reproductive, and hepatic systems (see 

Sections 6.3 to 6.5 and Sections 7.3 to 7.5 of the ISA, U.S. EPA, 2013). It has been proposed that 

lipid oxidation products resulting from reaction of O3 with lipids and/or cellular membranes in 

the ELF are responsible for systemic responses; however, it is not known whether they gain 

access to the circulation. Alternatively, release of diffusible mediators from the lung into the 

circulation may initiate or propagate inflammatory responses in the circulation or other organ 

systems. 

Responses to O3 exposure are variable within the population. Although studies have 

shown a large range of pulmonary function (i.e., spirometric) responses to O3 among healthy 

young adults, responses within an individual are relatively consistent over time. Other responses 

to O3 have also been characterized by a large degree of inter-individual variability. For example, 

a 3- to 20-fold difference among subjects in their studies in airways inflammation (i.e., 

neutrophilia influx) following O3 exposure has been reported (Schelegle et al., 1991 and Devlin 

et al., 1991, respectively). Reproducibility of an individual’s inflammatory response to O3

exposure in humans, measured as sputum neutrophilia, was demonstrated by Holz et al. (1999). 

Since individual inflammatory responses were relatively consistent across time, it was thought 

that inflammatory responsiveness reflected an intrinsic characteristic of the subject (Mudway and 

Kelly, 2000). While the basis for the observed inter-individual variability in responsiveness to O3

is not clear, section 5.4.2 of the ISA (U.S. EPA, 2013) discusses mechanisms that may underlie 

the variability in responses seen among individuals. Certain functional genetic polymorphisms, 

pre-existing conditions or diseases, nutritional status, lifestages, and co-exposures contribute to 

altered risk of O3-induced effects. 

 

5

 Metaplasia is the reversible replacement of one differentiated cell type with another mature differentiated cell type. 

The change from one type of cell to another may generally be a part of normal maturation process or caused by 

some sort of abnormal stimulus. In simplistic terms, it is as if the original cells are not robust enough to withstand 

the new environment, and so they change into another type more suited to the new environment. If the stimulus that 

caused metaplasia is removed or ceases, tissues return to their normal pattern of differentiation.
3-7 

Experimental evidence for such O3-induced changes contributes to our understanding of 

the biological plausibility of adverse O3-related health effects, including a range of respiratory 

effects as well as effects outside the respiratory system (e.g., cardiovascular effects) (U.S. EPA, 

2013, Chapters 6 and 7).

3.1.2 Nature of Effects 

 To what extent does the currently available scientific evidence alter or strengthen 

our conclusions from the last review regarding health effects attributable to O3

exposure in ambient air? Are previously identified uncertainties reduced or do 

important uncertainties remain? 

The health effects of ozone are described in detail in the assessment of the evidence 

available in this review which is largely consistent with conclusions of past Air Quality Criteria 

Documents (AQCD). In some categories of health effects, there is newly available evidence 

regarding some aspects of the effects described in the last review or that strengthens our 

conclusions regarding aspects of O3 toxicity on a particular physiological system (U.S. EPA, 

2013, Table 1-1). A sizeable number of studies on O3 health effects are newly available in this 

review and are critically assessed in the ISA as part of the full body of evidence. Based on this 

assessment, the ISA determined that a causal relationship6 exists between short-term exposure to 

O3 in ambient air7 and effects on the respiratory system and that a likely to be causal relationship8

exists between long-term exposure to O3 in ambient air and respiratory effects (U.S. EPA, 2013, 

pp. 1-6 to 1-7). As stated in the ISA, “[c]ollectively, a very large amount of evidence spanning 

several decades supports a relationship between exposure to O3 and a broad range of respiratory 

effects” (ISA, p. 1-6). Additionally, the ISA determined that the relationships between short-term 

exposures to O3 in ambient air and both total mortality and cardiovascular effects are likely to be 

causal, based on expanded evidence bases in the current review (U.S. EPA, 2013, pp. 1-7 to 1-8). 

In the ISA, EPA additionally determined that the currently available evidence for additional 

endpoints is suggestive of causal relationships between short-term (central nervous system 

effects) and long-term exposure (cardiovascular effects, central nervous system effects and total 

mortality) to ambient O3. Consistent with emphasis in past reviews on O3 health effects for which 

 

6

 Since the last O3 NAAQS review, the ISAs which have replaced CDs in documenting each review of the scientific 

evidence (or air quality criteria) employ a systematic framework for weighing the evidence and describing 

associated conclusions with regard to causality, using established descriptors, as summarized in section 1.3.1 above 

(U.S. EPA, 2013, Preamble). 

7

 In determining that a causal relationship exists for O3 with specific health effects, EPA has concluded that 

“[e]vidence is sufficient to conclude that there is a causal relationship with relevant pollutant exposures” (ISA, p. 

lxiv). 

8

 In determining a likely to be a causal relationship exists for O3 with specific health effects, EPA has concluded that 

“[e]vidence is sufficient to conclude that a causal relationship is likely to exist with relevant pollutant exposures, but 

important uncertainties remain” (ISA, p. lxiv). 
3-8 

the evidence is strongest, we place the greatest emphasis on studies of health effects that have 

been judged in the ISA to be caused by, or likely to be caused by, O3 exposures (U.S. EPA, 

2013, section 2.5.2). 

This section presents our consideration of the evidence for health effects attributable to 

O3 exposures, including respiratory morbidity and mortality effects attributable to short- and 

long-term exposures, and cardiovascular system effects (including mortality) and total mortality 

attributable to short-term exposures. We focus particularly on considering the extent to which the 

scientific evidence available in the current review has been strengthened since the last review, 

and the extent to which important uncertainties and limitations in the evidence from the last 

review have been addressed. In section 3.1.2.2, we then consider the extent to which the 

available evidence indicates health effects may be attributable to ambient O3 concentrations 

likely to be allowed by the current O3 NAAQS. In this section, we address the following specific 

question for each category of health effects considering the evidence available in the 2008 

review of the standard as well as evidence that has become available since then. The ISA 

summarizes the longstanding body of evidence for O3 respiratory effects as follows (U.S. EPA, 

2013, p. 1-6). 

The clearest evidence for health effects associated with exposure to O3 is provided 

by studies of respiratory effects. Collectively, a very large amount of evidence 

spanning several decades supports a relationship between exposure to O3 and a 

broad range of respiratory effects (see Section 6.2.9 and Section 7.2.8). The 

majority of this evidence is derived from studies investigating short-term 

exposures (i.e., hours to weeks) to O3, although animal toxicological studies and 

recent epidemiologic evidence demonstrate that long-term exposure (i.e., months 

to years) may also harm the respiratory system. 

The extensive body of evidence supporting a causal relationship between short-term O3

exposures and respiratory effects is discussed in detail in Chapter 6 of the ISA (U.S. EPA, 2013), 

while evidence for respiratory effects associated with long-term or repeated O3 exposures are 

discussed in chapter 7 of that document (U.S., EPA, 2013). 

3.1.2.1 Respiratory Effects – Short-term Exposures 

 To what extent does the currently available scientific evidence, including related 

uncertainties, strengthen or alter our understanding from the last review of respiratory 

effects attributable to short-term O3 exposures? 

The 2006 O3 AQCD concluded that there was clear, consistent evidence of a causal 

relationship between short-term O3 exposure and respiratory effects (U.S. EPA, 2006). This 

conclusion was substantiated by evidence from controlled human exposure and toxicological 

studies indicating a range of respiratory effects in response to short-term O3 exposures, including 

pulmonary function decrements and increases in respiratory symptoms, lung inflammation, lung 
3-9 

permeability, and airway hyperresponsiveness. Toxicological studies provided additional 

evidence for O3-induced impairment of host defenses. Combined, these findings from 

experimental studies provided support for epidemiologic evidence, in which short-term increases 

in ambient O3 concentration were consistently associated with decreases in lung function in 

populations with increased outdoor exposures, especially children with asthma and healthy 

children; increases in respiratory symptoms and asthma medication use in children with asthma; 

and increases in respiratory-related hospital admissions and asthma-related ED visits (U.S. EPA, 

2013, pp. 6-1 to 6-2). 

As discussed in detail in the ISA (U.S. EPA, 2013, section 6.2.9), studies evaluated since 

the completion of the 2006 O3 AQCD support and expand upon the strong body of evidence that, 

in the last review, indicated a causal relationship between short-term O3 exposures and 

respiratory health effects. Recent controlled human exposure studies conducted in young, healthy 

adults with moderate exertion have reported FEV1 decrements and pulmonary inflammation 

following prolonged exposures to O3 concentrations as low as 60 ppb, and respiratory symptoms 

following exposures to concentrations as low as 70 ppb.9

 Epidemiologic studies provide 

evidence that increases in ambient O3 exposures can result in lung function decrements, increases 

in respiratory symptoms, and pulmonary inflammation in children with asthma; increases in 

respiratory-related hospital admissions and emergency department visits; and increases in 

respiratory mortality. Some of these studies report such associations even for O3 concentrations 

at the low end of the distribution of daily concentrations. Recent epidemiologic studies report 

that associations with respiratory morbidity and mortality are stronger during the warm/summer 

months and remain robust after adjustment for copollutants. Recent toxicological studies 

reporting O3-induced inflammation, airway hyperresponsiveness, and impaired lung host defense 

continue to support the biological plausibility and modes of action for the O3-induced respiratory 

effects observed in the controlled human exposure and epidemiologic studies. Further support is 

provided by recent studies that found O3-associated increases in indicators of airway 

inflammation and oxidative stress in children with asthma (U.S. EPA, 2013, section 6.2.9). 

Together, epidemiologic and experimental studies support a continuum of respiratory effects 

associated with O3 exposure that can result in respiratory-related emergency department visits, 

hospital admissions, and/or mortality (U.S. EPA, 2013, section 6.2.9). 

Across respiratory endpoints, evidence indicates antioxidant capacity may modify the 

risk of respiratory morbidity associated with O3 exposure (U.S. EPA, 2013, section 6.2.9, p. 6-

161) (section 3.1.1, above). The potentially elevated risk of populations with diminished 

 

9

 Schelegle et al. (2009) reported a statistically significant increase in respiratory symptoms in healthy adults at a 

target O3 exposure concentration of 70 ppb. For this 70 ppb target, Schelegle et al. (2009) reported an actual 

exposure concentration, averaged over the study period, of 72 ppb. 
3-10 

antioxidant capacity and the reduced risk of populations with sufficient antioxidant capacity is 

supported by epidemiologic and controlled human exposure studies. Additional evidence 

characterizes O3-induced decreases in antioxidant levels as a key event in the mode of action for 

downstream effects. 

We describe key aspects of this evidence below with regard to lung function decrements; 

pulmonary inflammation, injury, and oxidative stress; airway hyperresponsiveness; respiratory 

symptoms and medication use; lung host defense; allergic and asthma-related responses; hospital 

admissions and emergency department visits; and respiratory mortality. 

Lung Function Decrements 

In the 2008 review, a large number of controlled human exposure studies reported O3-

induced lung function decrements in young, healthy adults engaged in intermittent, moderate 

exertion following 6.6 hour exposures to O3 concentrations at or above 80 ppb. Although two 

studies also reported effects following exposures to lower concentrations, an important 

uncertainty in the last review was the extent to which exposures to O3 concentrations below 80 

ppb result in lung function decrements. In addition, in the last review epidemiologic panel 

studies had reported O3-associated lung function decrements in a variety of different populations 

(e.g., children, outdoor workers) likely to experience increased exposures. In the current review, 

additional controlled human exposure studies are available that have evaluated exposures to O3

concentrations of 60 or 70 ppb. The available evidence from controlled human exposure and 

panel studies is assessed in detail in the ISA (U.S. EPA, 2013, section 6.2.1) and is summarized 

below. 

Controlled exposures to O3 concentrations that can be found in the ambient air can result 

in a number of lung function effects, including decreased inspiratory capacity; mild 

bronchoconstriction; and rapid, shallow breathing patterns during exercise. 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 second (FEV1) (U.S. EPA, 2013, section 6.2.1.1).10 Accumulating evidence 

indicates that such effects are mediated by activation of sensory nerves, resulting in the 

involuntary truncation of inspiration and a mild increase in airway obstruction due to 

bronchoconstriction (U.S. EPA, 2013, section 5.3.10). 

 

10 The controlled human exposure studies emphasized in this PA utilize only healthy adult subjects. In the near 

absence of controlled human exposure data for children, HREA estimates of lung function decrements are based on 

the assumption that children exhibit the same lung function responses following O3 exposures as healthy 18 year 

olds (U.S. EPA, 2014, sections 6.2.4 and 6.5). This assumption is justified in part by the findings of McDonnell et 

al. (1985), who reported that children 8-11 year old experienced FEV1 responses similar to those observed in adults 

18-35 years old. Thus, the conclusions about the occurrence of lung function decrements that follow generally apply 

to children as well as to adults. 
3-11 

Data from controlled human exposure studies indicate that increasing the duration of O3

exposures and increasing ventilation rates decreases the O3 exposure concentrations required to 

impair lung function. Ozone exposure concentrations well above those typically found in 

ambient air are required to impair lung function in healthy resting adults, while exposure to O3

concentrations at or below those in the ambient air have been reported to impair lung function in 

healthy adults exposed for longer durations while undergoing intermittent, moderate exertion 

(U.S. EPA, 2013, section 6.2.1.1). With repeated O3 exposures over several days, FEV1

responses become attenuated in both healthy adults and adults with mild asthma, though this 

attenuation of response is lost after about a week without exposure (U.S. EPA, 2013, section 

6.2.1.1; page 6-27). 

When considering controlled human exposures studies of O3-induced lung function 

decrements we evaluate both group mean changes in lung function and the interindividual 

variability in the magnitude of responses. An advantage of O3 controlled human exposure studies 

(i.e., compared to the epidemiologic panel studies discussed below) is that reported effects 

necessarily result from exposures to O3 itself.11 To the extent studies report statistically 

significant decrements in mean lung function following O3 exposures after controlling for other 

factors, we have more confidence that measured decrements are due to the O3 exposure itself, 

rather than to chance alone. As discussed below, group mean changes in lung function are often 

small, especially following exposures to relatively low O3 concentrations (e.g., 60 ppb). 

However, even when group mean decrements in lung function are small, some individuals could 

experience decrements that are “clinically meaningful” (Pellegrino et al., 2005; ATS, 1991) with 

respect to criteria for spirometric testing, and/or that could be considered “adverse” with respect 

to public health policy decisions (section 3.1.3 below). 

At the time of the last review, a number of controlled human exposure studies had 

reported lung function decrements in young, healthy adults following prolonged (6.6-hour) 

exposures while at moderate exertion to O3 concentrations at and above 80 ppb. In addition, 

there were two controlled human exposure studies by Adams (2002, 2006) that examined lung 

function effects following exposures to O3 concentrations of 60 ppb. The EPA’s analysis of the 

data from the Adams (2006) study reported a small but statistically significant O3-induced 

decrement in group mean FEV1 following exposures of young, healthy adults, while at moderate 

exertion, to 60 ppb O3, when compared with filtered air controls (Brown, 2008).12 Further 

 

11 The ISA notes that the use of filtered air responses as a control for the assessment of responses following O3 

exposure in controlled human exposure studies serves to eliminate alternative explanations other than O3 itself in 

causing the measured responses (U.S. EPA, 2013, section 6.2.1.1). 

12 Adams (2006) did not find effects on FEV1 at 60 ppb to be statistically significant. In an analysis of the Adams 

(2006) data, even after removal of potential outliers, Brown et al. (2008) found the average effect on FEV1 at 60 ppb 

to be small, but highly statistically significant (p < 0.002) using several common statistical tests. 
3-12 

examination of the post-exposure FEV1 data, and mean data for other time points and other 

concentrations, indicated that the temporal pattern of the response to 60 ppb O3 was generally 

consistent with the temporal patterns of responses to higher O3 concentrations in this and other 

studies. (75 FR 2950, January 19, 2010). This suggested a pattern of response following 

exposures to 60 ppb O3 that was consistent with a dose-response relationship, rather than random 

variability. See also State of Mississippi v. EPA, 744 F. 3d at 1347 (upholding EPA’s 

interpretation of the Adams studies). 

Figure 6-1 in the ISA summarizes the currently available evidence from multiple 

controlled human exposure studies evaluating group mean changes in FEV1 following prolonged 

O3 exposures (i.e., 6.6 hours) in young, healthy adults engaged in moderate levels of physical 

activity (U.S. EPA, 2013, section 6.2.1.1). With regard to the group mean changes reported in 

these studies, the ISA specifically notes the following (U.S. EPA, 2013, section 6.2.1.1, Figure 

6-1): 

1. Prolonged exposure to 40 ppb O3 results in a small decrease in group mean FEV1 that is 

not statistically different from responses following exposure to filtered air (Adams, 2002; 

Adams, 2006). 

2. Prolonged exposure to an average O3 concentration of 60 ppb results in group mean FEV1

decrements ranging from 1.8% to 3.6% (Adams 2002; Adams, 2006;13 Schelegle et al., 

2009;14 Kim et al., 2011). Based on data from multiple studies, the weighted average 

group mean decrement was 2.7%. In some analyses, these group mean decrements in 

lung function were statistically significant (Brown et al., 2008; Kim et al., 2011), while in 

other analyses they were not (Adams, 2006; Schelegle et al., 2009).15 

3. Prolonged exposure to an average O3 concentration of 70 ppb results in a statistically 

significant group mean decrement in FEV1 of about 6% (Schelegle et al., 2009).16 

4. Prolonged square-wave exposure to average O3 concentrations of 80 ppb, 100 ppb, or 120 

ppb O3 results in statistically significant group mean decrements in FEV1 ranging from 6 

to 8%, 8 to 14%, and 13 to 16%, respectively (Folinsbee et al., 1988; Horstman et al., 

1990; McDonnell et al., 1991; Adams, 2002; Adams, 2003; Adams, 2006). 

 

13 Adams (2006; 2002) both provide data for an additional group of 30 healthy subjects that were exposed via 

facemask to 60 ppb (square-wave) O3 for 6.6 hours with moderate exercise (Vሶ

୉ = 23 L/min per m2

 BSA). These 

subjects are described on page 133 of Adams (2006) and pages 747 and 761 of Adams (2002). The FEV1 decrement 

may be somewhat increased due to a target Vሶ

୉ of 23 L/min per m2

 BSA relative to other studies having the target Vሶ

୉

of 20 L/min per m2

 BSA. The facemask exposure is not expected to affect the FEV1 responses relative to a chamber 

exposure. 

14 Schelegle et al. (2009) reported an actual mean exposure concentration of 63 ppb for the target of 60 ppb. 

15 Adams (2006) did not find effects on FEV1 at 60 ppb to be statistically significant. In an analysis of the Adams 

(2006) data, Brown et al. (2008) addressed the more fundamental question of whether there were statistically 

significant differences in responses before and after the 6.6 hour exposure period and found the average effect on 

FEV1 at 60 ppb to be small, but highly statistically significant using several common statistical tests, even after 

removal of potential outliers. 

16 Schelegle et al. (2009) reported an actual mean exposure concentration of 72 ppb for the target of 70 ppb. 
3-13 

As illustrated in Figure 6-1 of the ISA, there is a smooth dose-response curve without 

evidence of a threshold for exposures between 40 and 120 ppb O3 (U.S. EPA, 2013, Figure 6-1). 

When these data are taken together, the ISA concludes that “mean FEV1 is clearly decreased by 

6.6-h exposures to 60 ppb O3 and higher concentrations in [healthy, young adult] subjects 

performing moderate exercise” (U.S. EPA, 2013, p. 6-9). 

 With respect to interindividual variability in lung function, in an individual with 

relatively “normal” lung function, with recognition of the technical and biological variability in 

measurements, within-day changes in FEV1 of ≥ 5% are clinically meaningful (Pellegrino et al., 

2005; ATS, 1991). The ISA (U.S. EPA, 2013, section 6.1.) focuses on individuals with >10% 

decrements in FEV1 for two reasons. A 10% FEV1 decrement is accepted by the American 

Thoracic Society (ATS) as an abnormal response and a reasonable criterion for assessing 

exercise-induced bronchoconstriction (Dryden et al., 2010; ATS, 2000). (U.S. EPA, 2013, 

section 6.2.1.1). Also, some individuals in the Schelegle et al. (2009) study experienced 5-10% 

FEV1 decrements following exposure to filtered air. 

In previous NAAQS reviews, the EPA has made judgments regarding the potential 

implications for individuals experiencing FEV1 decrements of varying degrees of severity.17 For 

people with lung disease, the EPA judged that moderate functional decrements (e.g., FEV1 

decrements > 10 percent but < 20 percent, lasting up to 24 hours) would likely interfere with 

normal activity for many individuals, and would likely result in more frequent use of medication 

(75 FR 2973, January 19, 2010). In previous reviews CASAC has endorsed these conclusions. In 

the context of standard setting, in the last review O3 review CASAC indicated that it is 

appropriate to focus on the lower end of the range of moderate functional responses (e.g., FEV1

decrements  10 percent) when estimating potentially adverse lung function decrements in 

people with lung disease, especially children with asthma (Henderson, 2006). More specifically, 

CASAC stated that “[a] 10% decrement in FEV1 can lead to respiratory symptoms, especially in 

individuals with pre-existing pulmonary or cardiac disease. For example, people with chronic 

obstructive pulmonary disease have decreased ventilatory reserve (i.e., decreased baseline FEV1) 

such that a ≥ 10% decrement could lead to moderate to severe respiratory symptoms” (Samet, 

2011). In this review, CASAC reiterated its support for this conclusion, stating that “[a]n FEV1 

decrement of ≥10% is a scientifically relevant surrogate for adverse health outcomes for people with 

asthma and lung disease” (Frey, 2014 p. 3). Therefore, in considering interindividual variability in 

 

17 Such judgments have been made for decrements in FEV1 as well as for increased airway responsiveness and 

symptomatic responses (e.g., cough, chest pain, wheeze). Ranges of pulmonary responses and their associated 

potential impacts are presented in Tables 3-2 and 3-3 of the Staff Paper (U.S. EPA, 2007). 
3-14 

O3-induced lung function decrements in the current review, we also focus on the extent to which 

individuals were reported to experience FEV1 decrements of 10% or greater. 

New studies (Schelegle et al., 2009; Kim et al., 2011) add to the previously available 

evidence for interindividual variability in the responses of healthy adults following exposures to 

O3. Following prolonged exposures to 80 ppb O3 while at moderate exertion, the proportion of 

healthy adults experiencing FEV1 decrements greater than 10% was 17% by Adams (2006), 26% 

by McDonnell (1996), and 29% by Schelegle et al. (2009). Following exposures to 60 ppb O3, 

that proportion was 20% by Adams (2002), 3% by Adams (2006), 16% by Schelegle et al. 

(2009), and 5% by Kim et al. (2011). Based on these studies, the weighted average proportion of 

young, healthy adults with >10% FEV1 decrements is 25% following exposure to 80 ppb O3 and 

10% following exposure to 60 ppb O3 (U.S. EPA, 2013, page 6-19).18 The ISA notes that 

responses within an individual tend to be reproducible over a period of several months, 

indicating that interindividual differences reflect differences in intrinsic responsiveness. Given 

this, the ISA concludes that “a considerable fraction” of healthy individuals experience clinically 

meaningful decrements in lung function when exposed for 6.6 hours to 60 ppb O3 during quasi 

continuous, moderate exertion (U.S. EPA, 2013, section 6.2.1.1, p. 6-20). 

As discussed above (Figure 3-1) and in the ISA (U.S EPA, 2013, Section 5.3.2), 

secondary oxidation products formed following O3 exposures can activate neural reflexes leading 

to decreased lung function. Two new quantitative models, discussed in section 6.2.1.1 of the ISA 

(U.S. EPA, 2013, p. 6-15), included mathematical approaches to simulate the protective effect of 

antioxidants in the ELF at lower ambient O3 concentrations, and include a threshold below which 

changes in lung function do not occur (McDonnell et al., 2012; Schelegle et al., 2012). 

McDonnell et al. (2012) and Schelegle et al. (2012) developed models using data on O3

exposure concentrations, ventilation rates, duration of exposures, and lung function responses 

from a number of controlled human exposure studies. The McDonnell et al. (2012) and Schelegle 

et al. (2012) studies analyzed large datasets to fit compartmental models that included the 

concept of a dose of onset in lung function response or a response threshold based upon the 

inhaled O3 dose. The first compartment in the McDonnell et al. (2012) model considers the level 

of oxidant stress in response to O3 exposure to increase over time as a function of dose rate 

(C×Vሶ

୉) and decrease by clearance or metabolism over time. In the second compartment of the 

McDonnell model, once oxidant stress reaches a threshold level the decrement in FEV1 increases 

 

18 The ISA also notes that by considering responses uncorrected for filtered air exposures, during which lung 

function typically improves (which would increase the size of the change, pre-and post-exposure), 10% is an 

underestimate of the proportion of healthy individuals that are likely to experience clinically meaningful changes in 

lung function following exposure for 6.6 hours to 60 ppb O3 during intermittent moderate exertion (U.S. EPA, 2013, 

section 6.2.1.1). 
3-15 

as a sigmoid-shaped function. In the Schelegle et al. (2012) model, a first compartment acts as a 

reservoir in which oxidant stress builds up until the dose of onset, at which time it spills over into 

a second compartment. The second compartment is identical to the first compartment in 

McDonnell et al. (2012) model. The oxidant levels in the second compartment were multiplied 

by a responsiveness coefficient to predict FEV1 responses for the Schelegle et al. (2012) model. 

The McDonnell et al. (2012) model was fit to a large dataset consisting of the FEV1

responses of 741 young, healthy adults (18-35 years of age) from 23 individual controlled 

exposure studies. Concentrations across individual studies ranged from 40 ppb to 400 ppb, 

activity level ranged from rest to heavy exercise, duration of exposure was from 2 to 7.6 hours. 

The extension of the McDonnell et al. (2012) model to children and older adults is discussed in 

section 6.2.1 of the ISA (U.S. EPA, 2013). Schelegle et al. (2012) also analyzed a large dataset 

with substantial overlap to that used by McDonnell et al. (2012). The Schelegle et al. (2012) 

model was fit to the FEV1 responses of 220 young healthy adults (taken from a dataset of 704 

individuals) from 21 individual controlled exposure studies. The resulting empirical models can 

estimate the frequency distribution of individual lung function responses for any exposure 

scenario as well as summary measures of the distribution such as the mean or median response 

and the proportions of individuals with FEV1 decrements > 10%, 15%, and 20%. 

The predictions of the McDonnell and Schelegle models are consistent with the observed 

results from the individual studies of O3-induced FEV1 decrements. Specifically, the model 

developed by McDonnell et al. (2012) predicts that 9% of healthy exercising adults would 

experience FEV1 decrements greater than 10% following 6.6 hour exposure to 60 ppb O3, and 

that 22% would experience such decrements following exposure to 80 ppb O3 (U.S. EPA, 2013, 

p. 6-18 and Figure 6-3). The model developed by Schelegle et al. (2012) predicts that, for a 

prolonged (6.6 hours) O3 exposure with moderate, quasi continuous exercise, the average dose of 

onset for FEV1 decrement would be reached following 4 to 5 hours of exposure to 60 ppb, and 

following 3 to 4 hours of exposure to 80 ppb. However, 14% of the individuals had a dose of 

onset that was less than 40% of the average. Those individuals would reach their dose of onset 

following 1 to 2 hours of exposure to 50 to 80 ppb O3 (U.S. EPA, 2013, p. 6-16), which is 

consistent with the threshold FEV1 responses reported by McDonnell et al. (2012). 

Epidemiologic studies19 have consistently linked short-term increases in ambient O3

concentrations with lung function decrements in diverse populations and lifestages, including 

children attending summer camps, adults exercising or working outdoors, and groups with preexisting respiratory diseases such as asthmatic children (U.S. EPA, 2013, section 6.2.1.2). Some 

 

19 Unless otherwise specified, the epidemiologic studies discussed in this PA evaluate only adults. 
3-16 

of these studies reported ozone-associated lung function decrements accompanied by respiratory 

symptoms20 in asthmatic children (Just et al., 2002; Mortimer et al., 2002; Ross et al., 2002; 

Gielen et al., 1997; Romieu et al., 1997; Thurston et al., 1997; Romieu et al., 1996). In contrast, 

studies of children in the general population have reported similar O3-associated lung function 

decrements but without accompanying respiratory symptoms (Ward et al., 2002; Gold et al., 

1999; Linn et al., 1996) (U.S. EPA, 2013, section 6.2.1.2). 

Several epidemiologic panel studies reported that associations with lung function 

decrements persisted at relatively low ambient O3 concentrations. For outdoor recreation or 

exercise, associations were reported in analyses restricted to 1-hour average O3 concentrations 

less than 80 ppb (Spektor et al., 1988a; Spektor et al., 1988b), 60 ppb (Brunekreef et al., 1994; 

Spektor et al., 1988a), and 50 ppb (Brunekreef et al., 1994). Among outdoor workers, Brauer et 

al. (1996) found a robust association using daily 1-hour max O3 concentrations less than 40 ppb. 

Ulmer et al. (1997) found a robust association in schoolchildren using 30-minute maximum O3

concentrations less than 60 ppb. For 8-hour average O3 concentrations, associations with lung 

function decrements in children with asthma were found to persist at concentrations less than 

80 ppb in a U.S. multicity study (Mortimer et al., 2002) and less than 51 ppb in a study 

conducted in the Netherlands (Gielen et al., 1997). 

Epidemiologic panel studies investigating the effects of short-term exposure to O3

provided information on potential confounding by copollutants such as PM2.5, PM10, NO2, or 

SO2. These studies varied in how they evaluated confounding. Some studies of subjects 

exercising outdoors indicated that ambient concentrations of copollutants such as NO2, SO2, or 

acid aerosol were low, and thus not likely to confound associations observed for O3 (Hoppe et 

al., 2003; Brunekreef et al., 1994; Hoek et al., 1993). In other studies of children with increased 

outdoor exposures, O3 was consistently associated with decreases in lung function, whereas other 

pollutants such as PM2.5, sulfate, and acid aerosol individually showed variable associations 

across studies (Thurston et al., 1997; Castillejos et al., 1995; Berry et al., 1991; Avol et al., 1990; 

Spektor et al., 1988a). Studies that conducted copollutant modeling generally found 

O3-associated lung function decrements to be robust (i.e., most copollutant-adjusted effect 

estimates fell within the 95% CI of the single-pollutant effect estimates) (U.S. EPA, 2013, Figure 

6-10 and Table 6-14). Most O3 effect estimates for lung function were robust to adjustment for 

temperature, humidity, and copollutants such as PM2.5, PM10, NO2, or SO2. Although examined 

 

20 Reversible loss of lung function in combination with the presence of symptoms meets the ATS definition of 

adversity (ATS, 2000). 
3-17 

in only a few epidemiologic studies, O3 also remained associated with decreases in lung function 

with adjustment for pollen or acid aerosols (U.S. EPA, 2013, section 6.2.1.2). 

Several epidemiologic studies demonstrated the protective effects of vitamin E and 

vitamin C supplementation, and increased dietary antioxidant intake, on O3-induced lung 

function decrements (Romieu et al., 2002) (U.S. EPA, 2013, Figure 6-7 and Table 6-8).21 These 

results provide support for the new, quantitative models (McDonnell et al., 2012; Schelegle et 

al., 2012), discussed above, which make use of the concept of oxidant stress to estimate the 

occurrence of lung function decrements following exposures to relatively low O3 concentrations. 

In conclusion, new information from controlled human exposure studies considerably 

strengthens the evidence and reduces the uncertainties, relative to the evidence that was available 

at the time of the 2008 review, regarding the presence and magnitude of lung function 

decrements in healthy adults following prolonged exposures to O3 concentrations below 80 ppb. 

As discussed in Section 6.2.1.1 in the ISA (U.S. EPA, 2013, p. 6-12), there is information 

available from four separate studies that evaluated exposures to 60 ppb O3 (Kim et al., 2011; 

Schelegle et al., 2009; Adams 2002; 2006). Although not consistently statistically significant, 

group mean FEV1 decrements following exposures to 60 ppb O3 are consistent among these 

studies. Moreover, as is illustrated in Figure 6-1 of the ISA (U.S. EPA, 2013), the group mean 

FEV1 responses at 60 ppb fall on a smooth intake dose-response curve for exposures between 40 

and 120 ppb O3. These studies also indicate that, on average, 10% of young, healthy adults 

experience clinically meaningful decrements in lung function when exposed for 6.6 hours to 60 

ppb O3 during intermittent, moderate exertion. One recent study has also reported statistically 

significant decrements following exposures to 70 ppb O3 (Schelegle et al., 2009). Predictions 

from newly developed quantitative models, based on the concept that O3-induced oxidation 

results in lung function decrements, are consistent with these experimental results. Additionally, 

as discussed in more detail in section 3.1.4 below, epidemiologic studies continue to provide 

evidence of lung function decrements in people who are active outdoors, including people 

engaged in outdoor recreation or exercise, children, and outdoor workers, at low ambient O3

concentrations. While few new epidemiologic studies of O3-associated lung function decrements 

are available in this review, previously available studies have reported associations with 

decrements, including at relatively low ambient O3 concentrations. 

Pulmonary Inflammation, Injury, and Oxidative Stress 

Ozone exposures result in increased respiratory tract inflammation and epithelial 

permeability. Inflammation is a host response to injury, and the induction of inflammation is 

 

21 Evidence from controlled human exposure studies is mixed, suggesting that supplementation may be ineffective 

in the absence of antioxidant deficiency (U.S. EPA, 2013, p. 5-63). 
3-18 

evidence that injury has occurred. Oxidative stress has been shown to play a key role in initiating 

and sustaining O3-induced inflammation. Secondary oxidation products formed as a result of 

reactions between O3 and components of the ELF can increase the expression of molecules (i.e., 

cytokines, chemokines, and adhesion molecules) that can enhance airway epithelium 

permeability (U.S. EPA, 2013, Sections 5.3.3 and 5.3.4). As discussed in detail in the ISA (U.S. 

EPA, 2013, section 6.2.3), O3 exposures can initiate an acute inflammatory response throughout 

the respiratory tract that has been reported to persist for at least 18-24 hours after exposure. 

Inflammation induced by exposure of humans to O3 can have several potential outcomes: 

(1) inflammation induced by a single exposure (or several exposures over the course of a 

summer) can resolve entirely; (2) continued acute inflammation can evolve into a chronic 

inflammatory state; (3) continued inflammation can alter the structure and function of other 

pulmonary tissue, leading to diseases such as asthma; (4) inflammation can alter the body’s host 

defense response to inhaled microorganisms, particularly in potentially at-risk populations or 

lifestages such as the very young and old; and (5) inflammation can alter the lung’s response to 

other agents such as allergens or toxins (U.S. EPA, 2013, Section 6.2.3). Thus, lung injury and 

the resulting inflammation provide a mechanism by which O3 may cause other more serious 

morbidity effects (e.g., asthma exacerbations). 

In the last review, controlled human exposure studies reported O3-induced airway 

inflammation following exposures at or above 80 ppb. In the current review, the link between O3

exposures and airway inflammation and injury has been evaluated in additional controlled human 

exposure studies, as well as in recent epidemiologic studies. Controlled human exposure studies 

have generally been conducted in young, healthy adults or in adults with asthma using lavage 

(proximal airway and bronchoalveolar), bronchial biopsy, and more recently, induced sputum. 

These studies have evaluated one or more indicators of inflammation, including neutrophil22

(PMN) influx, markers of eosinophilic inflammation, increased permeability of the respiratory 

epithelium, and/or prevalence of proinflammatory molecules (U.S. EPA, 2013, section 6.2.3.1). 

Epidemiologic studies have generally evaluated associations between ambient O3 and markers of 

inflammation and/or oxidative stress, which plays a key role in initiating and sustaining 

inflammation (U.S. EPA, 2013, section 6.2.3.2). 

There is an extensive body of evidence from controlled human exposure studies 

indicating that short-term exposures to O3 can cause pulmonary inflammation. Previously 

 

22 Referred to as either neutrophils or polymorphonuclear neutrophils (or PMNs), these are the most abundant type 

of white blood cells in mammals. PMNs are recruited to the site of injury following trauma and are the hallmark of 

acute inflammation. The presence of 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. Neutrophilic inflammation of tissues indicates 

activation of the innate immune system and requires a complex series of events, that then are normally followed by 

processes that clear the evidence of acute inflammation. 
3-19 

available evidence indicated that O3 causes an inflammatory response in the lungs (U.S. EPA, 

1996). A single acute exposure (1-4 hours) of humans to moderate concentrations of O3 (200-

600 ppb) while exercising at moderate to heavy intensities resulted in a number of cellular and 

biochemical changes in the lung, including inflammation characterized by increased numbers of 

PMNs, increased permeability of the epithelial lining of the respiratory tract, cell damage, and 

production of proinflammatory molecules (i.e., cytokines and prostaglandins, U.S. EPA, 2006). 

A meta-analysis of 21 controlled human exposure studies (Mudway and Kelly, 2004) using 

varied experimental protocols (80-600 ppb O3 exposures; 1-6.6 hours exposure duration; light to 

heavy exercise; bronchoscopy at 0-24 hours post-O3 exposure) reported that PMN influx in 

healthy subjects is linearly associated with total O3 dose. Animal toxicological studies also 

provided evidence for increases in inflammation and permeability in rabbits at levels as low as 

100 ppb O3 (Section 2.5.3.1, ISA, U.S. EPA, 2013). 

Several studies, including one published since the last review (Alexis et al., 2010), have 

reported O3-induced increases in PMN influx and permeability following exposures at or above 

80 ppb (Alexis et al., 2010; Peden et al., 1997; Devlin et al., 1991), and eosinophilic 

inflammation following exposures at or above 160 ppb (Scannell et al., 1996; Peden et al., 1997; 

Hiltermann et al., 1999; Vagaggini et al., 2002). In addition, one recent controlled human 

exposure study has reported O3-induced PMN influx following exposures of healthy adults to O3

concentrations of 60 ppb (Kim et al., 2011), the lowest concentration at which inflammatory 

responses have been evaluated in human studies. 

As with FEV1 responses to O3, inflammatory responses to O3 are generally reproducible 

within individuals, with some individuals experiencing more severe O3-induced airway 

inflammation than indicated by group averages (Holz et al., 2005; Holz et al., 1999). Unlike O3-

induced decrements in lung function, which are attenuated following repeated exposures over 

several days (U.S. EPA, 2013, section 6.2.1.1), some markers of O3-induced inflammation and 

tissue damage remain elevated during repeated exposures, indicating ongoing damage to the 

respiratory system (U.S. EPA, 2013, section 6.2.3.1, p. 6-81). 

Most controlled human exposure studies have reported that asthmatics experience larger 

O3-induced inflammatory responses than non-asthmatics. Specifically, asthmatics exposed to 

200 ppb O3 for 4-6 hours with exercise show significantly more neutrophils in bronchoalveolar 

lavage fluid (BALF) than similarly exposed healthy individuals (Scannell et al., 1996; Basha et 

al., 1994). Bosson et al. (2003) reported significantly greater expression of a variety of proinflammatory cytokines in asthmatics, compared to healthy subjects, following exposure to 

200 ppb O3 for 2 hours. In addition, research available in the last review, combined with a recent 

study newly available in this review, indicates that pretreatment of asthmatics with 

corticosteroids can prevent the O3-induced inflammatory response in induced sputum, though 
3-20 

pretreatment did not prevent FEV1 decrements (Vagaggini et al., 2001; 2007). In contrast, 

Stenfors et al. (2002) did not detect a difference in the O3-induced increases in neutrophil 

numbers between 15 subjects with mild asthma and 15 healthy subjects by bronchial wash at the 

6 hours postexposure time point, although the neutrophil increase in the asthmatic group was on 

top of an elevated baseline. 

In people with allergic airway disease, including people with rhinitis and asthma, 

evidence available in the last review indicated that proinflammatory mediators also cause 

accumulation of eosinophils in the airways (Jorres et al., 1996; Peden et al.,1995 and 1997; 

Frampton et al., 1997; Hiltermann et al., 1999; Holz et al., 2002; Vagaggini et al., 2002). The 

eosinophil, which increases inflammation and allergic responses, is the cell most frequently 

associated with exacerbations of asthma (75 CFR 2969, January 19, 2010). 

Studies reporting inflammatory responses and markers of lung injury have clearly 

demonstrated that there is important variation in the responses of exposed subjects (75 FR 2953, 

January 19, 2010). Some individuals also appear to be intrinsically more susceptible to increased 

inflammatory responses from O3 exposure (Holz et al., 2005). In healthy adults exposed to each 

80 and 100 ppb O3, Devlin et al. (1991) observed group average increases in neutrophilic 

inflammation of 2.1- and 3.8-fold, respectively. However, there was a 20-fold range in 

inflammatory responses between individuals at both concentrations. Relative to an earlier, 

similar study conducted at 400 ppb (Koren et al., 1989), Devlin et al. (1991) noted that although 

some of the study population showed little or no increase in inflammatory and cellular injury 

indicators analyzed after exposures to lower levels of O3 (i.e., 80 and 100 ppb), others had 

changes that were as large as those seen when subjects were exposed to 400 ppb O3. The data 

suggest that as a whole the healthy population, on average, may have small inflammatory 

responses to near-ambient levels of O3, though there may be a substantial subpopulation that is 

very sensitive to low levels of O3. Devlin et al. (1991) expressed the view that ‘‘susceptible 

subpopulations such as the very young, elderly, and people with pulmonary impairment or 

disease may be even more affected.’’ 

A number of studies report that O3 exposures increase epithelial permeability. Increased 

BALF protein, suggesting O3-induced changes in epithelial permeability, has been reported at 

1 hour and 18 hours postexposure (Devlin et al., 1997; Balmes et al., 1996). A meta-analysis of 

results from 21 publications (Mudway and Kelly, 2004) for varied experimental protocols (80-

600 ppb O3; 1-6.6 hours duration; light to heavy exercise; bronchoscopy at 0-24 hours post-O3

exposure; healthy subjects), showed that increased BALF protein is associated with total inhaled 

O3 dose (i.e., the product of O3 concentration, exposure duration, and Vሶ

୉). As noted in the 2009 

PM ISA (U.S. EPA, 2009), it has been postulated that changes in permeability associated with 

acute inflammation may provide increased access of inhaled antigens, particles, and other 
3-21 

inhaled substances deposited on lung surfaces to the smooth muscle, interstitial cells, immune 

cells underlying the epithelium, and the blood (U.S. EPA, 2013, sections 5.3.4, 5.3.5). Animal 

toxicology studies have provided some support for this hypothesis (Adamson and Prieditis, 1995; 

Chen et al., 2006), though these studies did not specifically evaluate O3 exposures (U.S. EPA, 

2009). Because of this potentially increased access, it has been postulated that increases in 

epithelial permeability following O3 exposure might lead to increases in airway responsiveness 

to specific and nonspecific agents. In a recent study, Que et al. (2011) investigated this 

hypothesis in healthy young adults (83M, 55 F) exposed to 220 ppb O3 for 2.25 hours 

(alternating 15 min periods of rest and brisk treadmill walking). As has been observed for FEV1

responses, within-individual changes in permeability were correlated between sequential O3

exposures, indicating intrinsic differences among individuals in susceptibility to epithelial 

damage following O3 exposures. However, increases in epithelial permeability at 1 day post-O3

exposure were not correlated with with changes in airway responsiveness assessed 1 day post-O3

exposure. The authors concluded that changes in epithelial permeability is relatively constant 

over time in young healthy adults, although changes in permeability and AHR appear to be 

mediated by different physiologic pathways. 

 The limited epidemiologic evidence reviewed in the 2006 O3 AQCD (U.S. EPA, 2006) 

demonstrated an association between short-term increases in ambient O3 concentrations and 

airways inflammation in children (1-hour max O3 of approximately 100 ppb). In the 2006 O3

AQCD (U.S. EPA, 2006), there was limited evidence for increases in nasal lavage levels of 

inflammatory cell counts and molecules released by inflammatory cells (i.e., eosinophilic 

cationic protein, and myeloperoxidases). Since 2006, as a result of the development of less 

invasive methods, there has been a large increase in the number of studies assessing ambient O3-

associated changes in airway inflammation and oxidative stress, the types of biological samples 

collected, and the types of indicators. Most of these recent studies have evaluated biomarkers of 

inflammation or oxidative stress in exhaled breath, nasal lavage fluid, or induced sputum (U.S. 

EPA, 2013, section 6.2.3.2). These recent studies form a larger database to establish coherence 

with findings from controlled human exposure and animal studies that have measured the same 

or related biological markers. Additionally, results from these studies provide further biological 

plausibility for the associations observed between ambient O3 concentrations and respiratory 

symptoms and asthma exacerbations. 

A number of epidemiologic studies provide evidence that short-term increases in ambient 

O3 exposure increase pulmonary inflammation and oxidative stress in children, including those 

with asthma (Sienra-Monge et al., 2004; Barraza-Villarreal et al., 2008; Romieu et al., 2008; 
3-22 

Berhane et al., 2011). Multiple studies examined and found increases in exhaled NO (eNO)23

(Berhane et al., 2011; Khatri et al., 2009; Barraza-Villarreal et al., 2008). In some studies of 

subjects with asthma, increases in ambient O3 concentration at the same lag were associated with 

both increases in pulmonary inflammation and respiratory symptoms (Khatri et al., 2009; 

Barraza-Villarreal et al., 2008). Although more limited in number, epidemiologic studies also 

found associations with cytokines such as IL-6 or IL-8 (Barraza-Villarreal et al., 2008; SienraMonge et al., 2004), eosinophils (Khatri et al., 2009), antioxidants (Sienra-Monge et al., 2004), 

and indicators of oxidative stress (Romieu et al., 2008) (ISA, Section 6.2.3.2, U.S. EPA, 2013). 

Because associations with inflammation were attenuated with higher antioxidant intake the study 

by Sienra-Monge et al. (2004) provides additional evidence that inhaled O3 is likely to be an 

important source of reactive oxygen species in airways and/or may increase pulmonary 

inflammation via oxidative stress-mediated mechanisms among all age groups. Limitations in 

some recent studies have contributed to inconsistent results in adults (U.S. EPA, 2013, section 

6.2.3.2). 

Exposure to ambient O3 on multiple days can result in larger increases in pulmonary 

inflammation and oxidative stress, as discussed in section 6.2.3.2 of the ISA (U.S. EPA, 2013). 

In studies that examined multiple O3 lags, multiday averages of 8-hour maximum or 

8-hour average concentrations were associated with larger increases in pulmonary inflammation 

and oxidative stress (Berhane et al., 2011; Delfino et al., 2010a; Sienra-Monge et al., 2004), 

consistent with controlled human exposure (U.S. EPA, 2013, section 6.2.3.1) and animal studies 

(U.S. EPA, 2013, section 6.2.3.3) reporting that some markers of pulmonary inflammation 

remain elevated with O3 exposures repeated over multiple days. Evidence from animal 

toxicological studies also clearly indicates that O3 exposures result in damage and inflammation 

in the lung (ISA, Section 5.3, U.S. EPA, 2013). In the few studies that evaluated the potential for 

confounding, O3 effect estimates were not confounded by temperature or humidity, and were 

robust to adjustment for PM2.5 or PM10 (Barraza-Villarreal et al., 2008; Romieu et al., 2008; 

Sienra-Monge et al., 2004). 

In conclusion, a relatively small number of controlled human exposure studies evaluating 

O3-induced airway inflammation have become available since the last review. For purposes of 

reviewing the current O3 NAAQS, the most important of these recent studies reported a 

statistically significant increase in airway inflammation in healthy adults at moderate exertion 

following exposures to 60 ppb O3, the lowest concentration that has been evaluated for 

inflammation. In addition, a number of recent epidemiologic studies report O3-associated 

 

23 Exhaled NO has been shown to be a useful biomarker for airway inflammation in large population-based studies 

(ISA, U.S. EPA, 2013, Section 7.2.4). 
3-23 

increases in markers of pulmonary inflammation, particularly in children. Thus, recent studies 

continue to support the evidence for airway inflammation and injury that was available in 

previous reviews, with new evidence for such effects following exposures to lower 

concentrations than had been evaluated previously. 

Airway Hyperresponsiveness 

Airway hyperresponsiveness (AHR) refers to a condition in which the conducting 

airways undergo enhanced bronchoconstriction in response to a variety of stimuli. Airway 

hyperresponsiveness is an important consequence of exposure to ambient O3 because its presence 

reflects a change in airway smooth muscle reactivity, and indicates that the airways are 

predisposed to narrowing upon inhalation of a variety of ambient stimuli including specific 

triggers (i.e., allergens) and nonspecific triggers (e.g., SO2, and cold air). People with asthma are 

generally more sensitive to bronchoconstricting agents than those without asthma, and the use of 

an airway challenge to inhaled bronchoconstricting agents is a diagnostic test in asthma. 

Standards for airway responsiveness testing have been developed for the clinical laboratory 

(ATS, 2000), although variation in the methodology for administering the bronchoconstricting 

agent may affect the results (Cockcroft et al., 2005). There is a wide range of airway 

responsiveness in people without asthma, and responsiveness is influenced by a number of 

factors, including cigarette smoke, pollutant exposures, respiratory infections, occupational 

exposures, and respiratory irritants. Dietary antioxidants have been reported to attenuate O3-

induced bronchial hyperresponsiveness in people with asthma (Trenga et al., 2001). 

Evidence for airway hyperresponsiveness following O3 exposures is derived primarily 

from controlled human exposure and toxicological studies (U.S. EPA, 2013, section 6.2.2). 

Airway responsiveness is often quantified by measuring changes in pulmonary function 

following the inhalation of an aerosolized allergen or a nonspecific bronchoconstricting agent 

(e.g., methacholine), or following exposure to a bronchoconstricting stimulus such as cold air. In 

the last review, controlled human exposure studies of mostly adults (≥ 18 years of age) had 

shown that exposures to O3 concentrations at or above 80 ppb increase airway responsiveness, as 

indicated by a reduction in the concentration of specific (e.g., ragweed) and non-specific (e.g., 

methacholine) agents required to produce a given reduction in lung function (e.g., as measured 

by FEV1 or specific airway resistance) (U.S. EPA, 2013, section 6.2.2.1). This O3-induced AHR 

has been reported to be dose-dependent (Horstman et al., 1990). Animal toxicology studies have 

reported O3-induced airway hyperresponsiveness in a number of species, with some rat strains 

exhibiting hyperresponsiveness following 4-hour exposures to O3 concentrations as low as 50 

ppb (Depuydt et al., 1999). Since the last review, there have been relatively few new controlled 

human exposure and animal toxicology studies of O3 and airway hyperresponsiveness, and no 
3-24 

new studies have evaluated exposures to O3 concentrations at or below 80 ppb (U.S. EPA, 2013, 

section 6.2.2.1) 

Airway hyperresponsiveness is linked with the accumulation and/or activation of 

eosinophils in the airways of asthmatics, which is followed by production of mucus and a latephase asthmatic response (75 FR 2970, January 19, 2010). In a study of 16 intermittent 

asthmatics, Hiltermann et al. (1999) found that there was a significant inverse correlation 

between the O3-induced change in the percentage of eosinophils in induced sputum and the 

concentration of methacholine causing a 20% decrease in FEV1. Hiltermann et al. (1999) 

concluded that the results point to the role of eosinophils in O3-induced airway 

hyperresponsiveness. Increases in O3-induced nonspecific airway responsiveness incidence and 

duration could have important clinical implications for children and adults with asthma, such as 

exacerbations of their disease. 

Airway hyperresponsiveness after O3 exposure appears to resolve more slowly than 

changes in FEV1 or respiratory symptoms (Folinsbee and Hazucha, 2000). Studies suggest that 

O3-induced AHR usually resolves 18 to 24 hours after exposure, but may persist in some 

individuals for longer periods (Folinsbee and Hazucha, 1989). Furthermore, in studies of 

repeated exposure to O3, changes in AHR tend to be somewhat less susceptible to attenuation 

with consecutive exposures than changes in FEV1 (Gong et al., 1997; Folinsbee et al., 1994; 

Kulle et al., 1982; Dimeo et al., 1981) (U.S. EPA, 2013, section 6.2.2). In animal studies a 3-day 

continuous exposure resulted in attenuation of O3-induced airway hyperresponsiveness (Johnston 

et al., 2005) while repeated exposures for 2 hours per day over 10 days did not (Chhabra et al., 

2010), suggesting that attenuation could be lost when repeated exposures are interspersed with 

periods of rest (U.S. EPA, 2013, section 6.2.2.2). 

Increases in airway responsiveness do not appear to be strongly associated with 

decrements in lung function or increases in symptoms (Aris et al., 1995). Recently, Que et al. 

(2011) assessed methacholine responsiveness in healthy young adults (83M, 55 F) one day after 

exposure to 220 ppb O3 and filtered air for 2.25 hours (alternating 15 minute periods of rest and 

brisk treadmill walking). Increases in airways responsiveness at 1 day post-O3 exposure were not 

correlated with FEV1 responses immediately following the O3 exposure or with changes in 

epithelial permeability assessed 1-day post-O3 exposure. This indicates that airway hyperresponsiveness also appears to be mediated by a differing physiologic pathway. 

As mentioned above, in addition to human subjects a number of species, including 

nonhuman primates, dogs, cats, rabbits, and rodents, have been used to examine the effect of O3

exposure on airway hyperresponsiveness (see Table 6-14, (U.S. EPA, 1996) of the 1996 O3

AQCD and Annex Table AX5-12 on page AX5-36 (U.S. EPA, 2006) of the 2006 O3 AQCD). A 

body of animal toxicology studies, including some recent studies conducted since the last review, 
3-25 

provides support for the O3-induced AHR reported in humans (U.S. EPA, 2013, section 6.2.2.2). 

Although most of these studies evaluated O3 concentrations above those typically found in 

ambient air in cities in the United States (i.e., most studies evaluated O3 concentrations of 100 

ppb or greater), one study reported that a very low exposure concentration (50 ppb for 4 hours) 

induced AHR in some rat strains (Depuydt et al., 1999). Additional recent rodent studies 

reported O3-induced AHR following exposures to O3 concentrations from 100 to 500 ppb 

(Johnston et al., 2005; Chhabra et al., 2010; Larsen et al., 2010). In characterizing the relevance 

of these exposure concentrations, the ISA noted that a study using radiolabeled O3 suggests that 

even very high O3 exposure concentrations in rodents could be equivalent to much lower 

exposure concentrations in humans. Specifically, a 2000 ppb (2 ppm) O3 exposure concentration 

in resting rats was reported to be roughly equivalent to a 400 ppb exposure concentration in 

exercising humans (Hatch et al., 1994). Given this relationship, the ISA noted that animal data 

obtained in resting conditions could underestimate the risk of effects for humans (U.S. EPA, 

2013, section 2.4, p. 2-14). 

The 2006 AQCD (U.S. EPA, 2006, p. 6-34) concluded that spirometric responses to O3

are independent of inflammatory responses and markers of epithelial injury or integrity (Balmes 

et al., 1996; Blomberg et al., 1999; Torres et al., 1997). Significant inflammatory responses to O3

exposures that did not elicit significant spirometric responses have been reported (Holz et al., 

2005). A recent study (Que et al., 2011) indicates that airway hyper-responsiveness also appears 

to be mediated by a differing physiologic pathway. These results from controlled human 

exposure studies indicate that sub-populations of healthy study subjects consistently experience 

larger than average lung function decrements, greater than average inflammatory responses and 

pulmonary injury as expressed by increased epithelial permeability, and greater than average 

airway responsiveness, and that these effects are mediated by apparently different physiologic 

pathways. Except for lung function decrements, we do not have the concentration- or exposureresponse function information about the other, potentially more sensitive,24 clinical endpoints 

(i.e., inflammation, increased epithelial permeability, airway hyperresponsiveness) that would 

allow us to quantitatively estimate the size of the population affected and the magnitude of their 

responses. Moreover, some uncertainties about the exact physiological pathways underlying 

these endpoints prevents us from knowing whether the exaggerated responses are distributed in 

sub-populations evenly across the population, or may be clustered with more than one type of 

exaggerated response in particular sub-populations, or both. 

 

24 CASAC noted that “...[W]hile measures of FEV1 are quantitative and readily obtainable in humans, they are not 

the only measures — and perhaps not the most sensitive measures — of the adverse health effects induced by ozone 

exposure.” (Henderson, 2006). 
3-26 

In summary, a strong body of controlled human exposure and animal toxicological 

studies, most of which were available in the last review of the O3 NAAQS, report O3-induced 

airway hyperresponsiveness after either acute or repeated exposures (U.S. EPA, 2013, section 

6.2.2.2). People with asthma often exhibit increased airway responsiveness at baseline relative to 

healthy controls, and they can experience further increases in responsiveness following 

exposures to O3. Studies reporting increased airway responsiveness after O3 exposure contribute 

to a plausible link between ambient O3 exposures and increased respiratory symptoms in 

asthmatics, and increased hospital admissions and emergency department visits for asthma (U.S. 

EPA, 2013, section 6.2.2.2). 

Respiratory Symptoms and Medication Use 

 Because respiratory symptoms are associated with adverse outcomes such as limitations 

in activity, and are the primary reason for people with asthma to use quick relief medication and 

seek medical care, studies evaluating the link between O3 exposures and such symptoms allow a 

more direct characterization of the clinical and public health significance of ambient O3 exposure 

than measures of lung function decrements and pulmonary inflammation. Controlled human 

exposure and toxicological studies have described modes of action through which short-term O3

exposures may increase respiratory symptoms by demonstrating O3-induced airway 

hyperresponsiveness (U.S. EPA, 2013, section 6.2.2) and pulmonary inflammation (U.S. EPA, 

2013, section 6.2.3). 

The link between subjective respiratory symptoms and O3 exposures has been evaluated 

in both controlled human exposure and epidemiologic studies, and the link with medication use 

has been evaluated in epidemiologic studies. In the last review, several controlled human 

exposure studies reported respiratory symptoms following exposures to O3 concentrations at or 

above 80 ppb. In addition, one study reported such symptoms following exposures to 60 ppb O3, 

though the increase was not statistically different from filtered air controls. Epidemiologic 

studies reported associations between ambient O3 and respiratory symptoms and medication use 

in a variety of locations and populations, including asthmatic children living in U.S. cities. In the 

current review, additional controlled human exposure studies have evaluated respiratory 

symptoms following exposures to O3 concentrations below 80 ppb and recent epidemiologic 

studies have evaluated associations with respiratory symptoms and medication use (U.S. EPA, 

2013, sections 6.2.1, 6.2.4). 

In controlled human exposure studies available in the last review as well as newly 

available studies, statistically significant increases in respiratory symptoms have been 

consistently reported in healthy adult volunteers engaged in intermittent, moderate exertion 

following 6.6 hour exposures to average O3 concentrations at or above 80 ppb (Adams, 2003; 

Adams, 2006; Schelegle et al., 2009). Such symptoms have been reported to increase with 
3-27 

increasing O3 exposure concentrations, duration of exposure, and activity level (McDonnell et 

al., 1999). For example, in a study available during the last review, Adams (2006) reported an 

increase in respiratory symptoms in healthy adults during a 6.6 hour exposure protocol with an 

average O3 exposure concentration of 60 ppb. This increase was significantly different from 

initial respiratory symptoms, but not from filtered air controls. Two recent controlled human 

exposure studies that have become available since the last review did not report statistically 

significant increases in respiratory symptoms following exposures of healthy adults to 60 ppb O3

(Schelegle et al., 2009; Kim et al., 2011). A recent study by Schelegle et al. (2009) did report a 

statistically significant increase in respiratory symptoms in healthy adults following 6.6 hour 

exposures to an average O3 concentration of 70 ppb. The findings for O3-induced respiratory 

symptoms in controlled human exposure studies, and the evidence integrated across disciplines 

describing underlying modes of action, provide biological plausibility for epidemiologic 

associations observed between short-term increases in ambient O3 concentration and increases in 

respiratory symptoms (U.S. EPA, 2013, section 6.2.4). 

In epidemiologic studies of respiratory symptoms, data typically are collected by having 

subjects (or their parents) record symptoms and medication use in a diary without direct 

supervision by study staff. Several limitations of symptom reports are well recognized, as 

described in the ISA (U.S. EPA, 2013, section 6.2.4). Nonetheless, symptom diaries remain a 

convenient tool to collect individual-level data from a large number of subjects and allow 

modeling of associations between daily changes in O3 concentration and daily changes in 

respiratory morbidity over multiple weeks or months. Importantly, many of the limitations in 

these studies are sources of random measurement error that can bias effect estimates to the null 

or increase the uncertainty around effect estimates (U.S. EPA, 2013, Section 6.2.4). Because 

respiratory symptoms are associated with limitations in activity and daily function and are the 

primary reason for using medication and seeking medical care, the evidence is directly coherent 

with the associations consistently observed between increases in ambient O3 concentration and 

increases in asthma emergency department visits, discussed below (U.S. EPA, 2013, Section 

6.2.4). 

Most epidemiologic studies of O3 and respiratory symptoms and medication use have 

been conducted in children and/or adults with asthma, with fewer studies, and less consistent 

results, in non-asthmatic populations (U.S. EPA, 2013, section 6.2.4). The 2006 AQCD (U.S. 

EPA, 2006, U.S. EPA, 2013, section 6.2.4) concluded that the collective body of epidemiologic 

evidence indicated that short-term increases in ambient O3 concentrations are associated with 

increases in respiratory symptoms in children with asthma. A large body of single-city and 

single-region studies of asthmatic children provides consistent evidence for associations between 
3-28 

short-term increases in ambient O3 concentrations and increased respiratory symptoms and 

asthma medication use in children with asthma (U.S. EPA, 2013, Figure 6-12, Table 6-20). 

Methodological differences among studies make comparisons across recent multicity 

studies of respiratory symptoms difficult. Because of fewer person-days of data (Schildcrout et 

al., 2006) or examination of 19-day averages of ambient O3 concentrations (O'Connor et al., 

2008), the ISA did not give greater weight to results from recent multicity studies than results 

from single-city studies (U.S. EPA, 2013, section 6.2.4.5). While evidence from the few 

available U.S. multicity studies is less consistent (O'Connor et al., 2008; Schildcrout et al., 2006; 

Mortimer et al., 2002), the overall body of epidemiologic evidence with respect to the 

association betweeen exposure to O3 and respiratory symptoms in asthmatic children remains 

compelling (U.S. EPA, 2013, section 6.2.4.1). Findings from a small body of studies indicate that 

O3 is also associated with increased respiratory symptoms in adults with asthma (Khatri et al., 

2009; Feo Brito et al., 2007; Ross et al., 2002) (U.S. EPA, 2013, section 6.2.4.2). 

Available evidence indicates that O3-associated increases in respiratory symptoms are not 

confounded by temperature, pollen, or copollutants (primarily PM) (U.S. EPA, 2013, section 

6.2.4.5; Table 6-25; Romieu et al., 1996; Romieu et al., 1997; Thurston et al., 1997; Gent et al., 

2003). However, identifying the independent effects of O3 in some studies was complicated due 

to the high correlations observed between O3 and PM or different lags and averaging times 

examined for copollutants. Nonetheless, the ISA noted that the robustness of associations in 

some studies of individuals with asthma, combined with findings from controlled human 

exposure studies for the direct effects of O3 exposure, provide substantial evidence supporting 

the independent effects of short-term ambient O3 exposure on respiratory symptoms (U.S. EPA, 

2013, section 6.2.4.5). 

Epidemiologic studies of medication use have reported associations with 

1-hour maximum O3 concentrations and with multiday average O3 concentrations (Romieu et al., 

2006; Just et al., 2002). Some studies reported O3 associations for both respiratory symptoms and 

asthma medication use (Escamilla-Nuñez et al., 2008; Romieu et al., 2006; Schildcrout et al., 

2006; Jalaludin et al., 2004; Romieu et al., 1997; Thurston et al., 1997) while others reported 

associations for either respiratory symptoms or medication use (Romieu et al., 1996; Rabinovitch 

et al., 2004; Just et al., 2002; Ostro et al., 2001). 

In summary, both controlled human exposure and epidemiologic studies have reported 

respiratory symptoms attributable to short-term O3 exposures. In the last review, the majority of 

the evidence from controlled human exposure studies in young, healthy adults was for symptoms 

following exposures to O3 concentrations at or above 80 ppb. Although studies that have become 

available since the last review have not reported respiratory symptoms in young, healthy adults 

following exposures with moderate exertion to 60 ppb, one recent study has reported increased 
3-29 

symptoms in young, healthy adults while at moderate exertion following exposures to O3

concentrations as low as 70 ppb.25 As was concluded in the 2006 O3 AQCD (U.S. EPA, 2006, 

1996), the collective body of epidemiologic evidence indicates that short-term increases in 

ambient O3 concentration are associated with increases in respiratory symptoms in children with 

asthma (U.S. EPA, 2013, section 6.2.4). Recent studies of respiratory symptoms and medication 

use, primarily in asthmatic children, add to this evidence. In a smaller body of studies, increases 

in ambient O3 concentration were associated with increases in respiratory symptoms in adults 

with asthma. 

Lung Host Defense 

The mammalian respiratory tract has a number of closely integrated defense mechanisms 

that, when functioning normally, provide protection from the potential health effects of 

exposures to a wide variety of inhaled particles and microbes. These defense mechanisms 

include mucociliary clearance, alveolobronchiolar transport mechanism, alveolar macrophages26, 

and adaptive immunity27 (U.S. EPA, 2013, section 6.2.5). The previous O3 AQCD (U.S. EPA, 

2006) concluded that animal toxicological studies provided evidence that acute exposure to O3

concentrations as low as 100 to 500 ppb can increase susceptibility to infectious diseases due to 

modulation of these lung host defenses. This conclusion was based in large part on animal 

studies of alveolar macrophage functioning and mucociliary clearance (U.S. EPA, 2013, section 

6.2.5). 

With regard to alveolar macrophage functioning, the previous O3 AQCD (U.S. EPA, 

2006) concluded that short periods of O3 exposure can cause a reduction in the number of free 

alveolar macrophages available for pulmonary defense, and that these alveolar macrophages are 

more fragile, less able to engulf particles (i.e., phagocytic), and exhibit decreased lysosomal28

enzyme activities (U.S. EPA, 2013, section 6.2.5). These conclusions were based largely on 

studies conducted in animals exposed for several hours up to several weeks to O3 concentrations 

from 100 to 250 ppb (Hurst et al., 1970; Driscoll et al., 1987; Cohen et al., 2002). Consistent 

with the animal evidence, a controlled human exposure study available in the last review had 

reported decrements in the ability of alveolar macrophages to phagocytize yeast following 

exposures of healthy volunteers to O3 concentrations of 80 and 100 ppb for 6.6-hour during 

 

25As noted above, for the 70 ppb exposure concentration Schelegle et al. (2009) reported that the actual mean 

exposure concentration was 72 ppb. 

26 Phagocytic white blood cells within the alveoli of the lungs that ingest inhaled particles. 27 The adaptive immune system, is also known as the acquired immune system. Acquired immunity creates 

immunological memory after an initial response to a specific pathogen, leading to an enhanced response to 

subsequent encounters with that same pathogen. 

28 Lysosomes are cellular organelles that contain acid hydrolase enzymes that break down waste materials and 

cellular debris. 
3-30 

moderate exercise (Devlin et al., 1991). Integrating the animal study results with human 

exposure evidence available in the 1996 Criteria Document, the 2006 Criteria Document 

concluded that available evidence indicates that short-term O3 exposures have the potential to 

impair host defenses in humans, primarily by interfering with alveolar macrophage function. Any 

impairment in alveolar 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 (EPA, 

2006, p. 8–26). 

With regard to mucociliary clearance, in the last review a number of studies conducted in 

different animal species had reported morphological damage to the cells of the tracheobronchial 

tree from acute and sub-chronic exposure to O3 concentrations at or above 200 ppb. The cilia 

were either completely absent or had become noticeably shorter or blunt. In general, functional 

studies of mucociliary transport had observed a delay in particle clearance soon after acute 

exposure, with decreased clearance more evident at higher doses (1 ppm) (U.S. EPA, 2013, 

section 6.2.5.1). 

Alveolobronchiolar transport mechanisms refers to the transport of particles deposited in 

the deep lung (alveoli) which may be removed either up through the respiratory tract (bronchi) 

by alveolobronchiolar transport or through the lymphatic system. The pivotal mechanism of 

alveolobronchiolar transport involves the movement of alveolar macrophages with ingested 

particles to the bottom of the conducting airways. These airways are lined with ciliated epithelial 

cells and cells that produce mucous, which surrounds the macrophages. The ciliated epithelial 

cells move the mucous packets up the resiratory tract, hence the term “mucociliary escalator.” 

Although some studies show reduced tracheobronchial clearance after O3 exposure, alveolar 

clearance of deposited material is accelerated, presumably due to macrophage influx, which in 

itself can be damaging. 

With regard to adaptive immunity, a limited number of epidemiologic studies have 

examined associations between O3 exposure and hospital admissions or ED visits for respiratory 

infection, pneumonia, or influenza. Results have been mixed, and in some cases conflicting (U.S. 

EPA, 2013, Sections 6.2.7.2 and 6.2.7.3). With the exception of influenza, it is difficult to 

ascertain whether cases of respiratory infection or pneumonia are of viral or bacterial etiology. 

A recent study that examined the association between O3 exposure and respiratory hospital 

admissions in response to an increase in influenza intensity did observe an increase in respiratory 

hospital admissions (Wong et al., 2009), but information from toxicological studies of O3 and 

viral infections is ambiguous. 

In summary, relatively few studies conducted since the last review have evaluated the 

effects of O3 exposures on lung host defense. When the available evidence is taken as a whole, 
3-31 

the ISA concludes that acute O3 exposures impair the host defense capability of animals, 

primarily by depressing alveolar macrophage function and perhaps also by decreasing 

mucociliary clearance of inhaled particles and microorganisms. Coupled with limited evidence 

from controlled human exposure studies, this suggests that humans exposed to O3 could be 

predisposed to bacterial infections in the lower respiratory tract (EPA, 2013, section 6.2.5.5). 

The seriousness of such infections may depend on how quickly bacteria develop virulence 

factors and how rapidly PMNs are mobilized to compensate for the deficit in alveolar 

macrophage function. 

Allergic and Asthma-Related Responses 

Effects resulting from combined exposures to O3 and allergens have been studied in a 

variety of animal species, generally as models of experimental asthma. Pulmonary function and 

AHR in animal models of asthma are discussed in detail in Section 6.2.1.3 and Section 6.2.2.2, 

respectively, in the ISA (U.S. EPA, 2013). Studies of allergic and asthma-related responses are 

discussed in detail in sections 5.3.6 and 6.2.6 of the ISA (U.S. EPA, 2013). 

Evidence available in the last review indicates that O3 exposure skews immune responses 

toward an allergic phenotype. For example, Gershwin et al. (1981) reported that O3 (800 and 500 

ppb for 4 days) exposure caused a 34-fold increase in the number of IgE (allergic antibody)-

containing cells in the lungs of mice. In general, the number of IgE-containing cells correlated 

positively with levels of anaphylactic sensitivity. In humans, allergic rhinoconjunctivitis 

symptoms are associated with increases in ambient O3 concentrations (Riediker et al., 2001). 

Controlled human exposure studies have observed O3-induced changes indicating allergic 

skewing. Airway eosinophils, which are white blood cells that participate in allergic disease and 

inflammation, were observed to increase in volunteers with atopy29 and mild asthma (Peden et 

al., 1997). In a more recent study, expression of IL-5, a cytokine involved in eosinophil 

recruitment and activation, was increased in subjects with atopy but not in healthy subjects 

(Hernandez et al., 2010). Epidemiologic studies describe associations between eosinophils in 

both short- (U.S. EPA, 2013, Section 6.2.3.2) and long-term (U.S. EPA, 2013, Section 7.2.5) O3

exposure, as do chronic exposure studies in non-human primates. Collectively, findings from 

these studies suggest that O3 can induce or enhance certain components of allergic inflammation 

in individuals with allergy or allergic asthma. 

Evidence available in the last review indicates that ozone may also increase AHR to 

specific allergen triggers (75 FR 2970, January 19, 2010). Two studies (Jörres et al., 1996; Holz 

et al., 2002) observed increased airway responsiveness to O3 exposure with bronchial allergen 

 

29 Atopy is a predisposition toward developing certain allergic hypersensitivity reactions. A person with atopy 

typically presents with one or more of the following: eczema (atopic dermatitis), allergic rhinitis (hay fever), allergic 

conjunctivitis, or allergic asthma. 
3-32 

challenge in subjects with preexisting allergic airway disease. Ozone-induced exacerbation of 

airway responsiveness persists longer and attenuates more slowly than O3-induced lung function 

decrements and respiratory symptom responses and can have important clinical implications for 

asthmatics. 

Animal toxicology studies indicate that O3 enhances inflammatory and allergic responses 

to allergen challenge in sensitized animals. In addition to exacerbating existing allergic 

responses, toxicology studies indicate that O3 can also act as an adjuvant to produce sensitization 

in the respiratory tract. Along with its pro-allergic effects (inducing or enhancing certain 

components of allergic inflammation in individuals with allergy or allergic asthma), O3 could 

also make airborne allergens more allergenic. When combined with NO2, O3 has been shown to 

enhance nitration of common protein allergens, which may increase their allergenicity Franze et 

al. (2005). 

Hospital Admissions and Emergency Department Visits 

The 2006 O3 AQCD evaluated numerous studies of respiratory-related emergency 

department visits and hospital admissions. These were primarily time-series studies conducted in 

the U.S., Canada, Europe, South America, Australia, and Asia. Based on such studies, the 2006 

O3 AQCD concluded that “the overall evidence supports a causal relationship between acute 

ambient O3 exposures and increased respiratory morbidity resulting in increased ED visits and 

[hospital admissions] during the warm season30” (U.S. EPA, 2006). This conclusion was 

“strongly supported by the human clinical, animal toxicologic[al], and epidemiologic evidence 

for [O3-induced] lung function decrements, increased respiratory symptoms, airway 

inflammation, and airway hyperreactivity” (U.S. EPA, 2006). 

The results of recent studies largely support the conclusions of the 2006 O3 AQCD (U.S. 

EPA, 2013, section 6.2.7). Since the completion of the 2006 O3 AQCD, relatively fewer studies 

conducted in the U.S., Canada, and Europe have evaluated associations between short-term O3

concentrations and respiratory hospital admissions and emergency department visits, with a 

growing number of studies conducted in Asia. This epidemiologic evidence is summarized in 

Appendix 3A and discussed in detail in the ISA (U.S. EPA, 2013, section 6.2.7). 

In considering this body of evidence, the ISA focused primarily on multicity studies 

because they examine associations with respiratory-related hospital admissions and emergency 

department visits over large geographic areas using consistent statistical methodologies (U.S. 

EPA, 2013, section 6.2.7.1). The ISA also focused on single-city studies that encompassed a 

large number of daily hospital admissions or emergency department visits, included long study-

 

30Epidemiologic associations for O3 are more robust during the warm season than during cooler months (e.g., 

smaller measurement error, less potential confounding by copollutants). Rationale for focusing on warm season 

epidemiologic studies for O3 can be found at 72 FR 37838-37840. 
3-33 

durations, were conducted in locations not represented by the larger studies, or examined 

population-specific characteristics that may increase the risk of O3-related health effects but were 

not evaluated in the larger studies (U.S. EPA, 2013, section 6.2.7.1). When examining the 

association between short-term O3 exposure and respiratory health effects that require medical 

attention, the ISA distinguishes between hospital admissions and emergency department visits 

because it is likely that a small percentage of respiratory emergency department visits will be 

admitted to the hospital; therefore, respiratory emergency department visits may represent 

potentially less serious, but more common outcomes (U.S. EPA, 2013, section 6.2.7.1). 

Several recent multicity studies (e.g., Cakmak et al., 2006; Dales et al., 2006) and a 

multi-continent study (Katsouyanni et al., 2009) report associations between short-term O3

concentrations and increased respiratory-related hospital admissions and emergency department 

visits. These multicity studies are supported by single-city studies also reporting consistent 

positive associations using different exposure assignment approaches (i.e., average of multiple 

monitors, single monitor, population-weighted average) and averaging times (i.e., 1-hour max 

and 8-hour max) (U.S. EPA, 2013, sections 6.2.7.1 to 6.2.7.5). When examining cause-specific 

respiratory outcomes, recent studies report positive associations with hospital admissions and 

emergency department visits for asthma (Strickland et al., 2010; Stieb et al., 2009) and COPD 

(Stieb et al., 2009; Medina-Ramon et al., 2006), with more limited evidence for pneumonia 

(Medina-Ramon et al., 2006; Zanobetti and Schwartz, 2006). In seasonal analyses (Figure 3-2 

below; U.S. EPA, 2013, Figure 6-19, Table 6-28), stronger associations were reported in the 

warm season or summer months (red circles), when O3 concentrations are higher, compared to 

the cold season (blue circles), particularly for asthma (Strickland et al., 2010; Ito et al., 2007) and 

COPD (Medina-Ramon et al., 2006).31 The available evidence indicates that children are at 

greatest risk for O3-induced respiratory effects (Silverman and Ito, 2010; Strickland et al., 2010; 

Mar and Koenig, 2009; Villeneuve et al., 2007; Dales et al., 2006). 

Although the collective evidence across studies indicates a mostly consistent positive 

association between O3 exposure and respiratory-related hospital admissions and ED visits, the 

magnitude of these associations may be underestimated due to behavioral modification in 

response to air quality forecasts (U.S. EPA, 2013, Section 4.6.6).

 

31 The study by Strickland et al. (2010) is discussed in more detail in section 3.1.4.2, below. 
3-34 

Note: Effect estimates are for a 20 ppb increase in 24-hour; 30 ppb increase in 8-hour max; and 40 ppb increase in 1-hour max O3 concentrations. HA=ho

ED=emergency department. Black=All-year analysis; Red=Summer only analysis; Blue=Winter only analysis. 

a Wheeze used as indicator of lower respiratory disease. b APHENA-Canada results standardized to approximate IQR of 5.1 ppb for 1-h max O3 concentrations. c Study included 8 cities; but of those 8, only 4 had O3 data. d non-ICU effect estimates. e The study did not specify the lag day of the summer season estimate. Figure 3-2. Percent increase in respiratory-related hospital admission and emergency department visits i

presented all-year and/or seasonal results.

Study

Wong et al. (2009) Cakmak et al. (2006) Dales et al. (2006) Orazzo et al. (2009)a Katsouyanni et al. (2009)

Darrowet al. (2009) Tolbertet al. (2007) Biggeri et al. (2005)c Katsouyanni et al. (2009)

Stieb et al. (2009) Villeneuve et al. (2007) Strickland et al. (2010) Silverman and Ito (2010)d Ito et al. (2007) Villeneuve et al. (2007) Mar and Koenig (2009) Strickland et al. (2010) Silverman and Ito (2010)d Mar and Koenig (2009) Ito et al. (2007) Villeneuve et al. (2007) Strickland et al. (2010)

Wong et al. (2009) Stieb et al. (2009) Yang et al. (2006) Medina‐Ramon et al. (2006) Stieb et al. (2009)e Medina‐Ramon et al. (2006)

Zanobetti and Schwartz (2006) Medina‐Ramon et al. (2006)

Location

Hong Kong 10 Canadian cities

11 Canadian cities

6 Italian cities

APHENA‐Europe APHENA‐U.S.

APHENA‐CanadaAPHENA‐Canada

AtlantaAtlanta

8 Italian cities

APHENA‐Europe APHENA‐U.S.

APHENA‐CanadaAPHENA‐Canada

7 Canadian Cities

Alberta, CAN Atlanta

New YorkNew York

Alberta, CAN Seattle, WA Atlanta

New York

Seattle, WA

New York

Alberta, CAN Atlanta

Hong Kong 7 Canadian Cities

Vancouver

36 U.S. cities

7 Canadian Cities

36 U.S. cities

36 U.S. cities

Boston

36 U.S. cities

36 U.S. cities

36 U.S. cities

Age

AllAll

0‐27 days 0‐2

65+65+65+65+

AllAllAll

65+65+65+65+

All

> 2

Children

AllAll

> 2

18+

Children

6‐18

<18

All

> 2

Children

AllAll

65+65+

All

65+65+65+65+65+65+

Respiratory

Asthma

COPD

Pneumonia

◄

Lag

0‐11.2

2

0‐60‐10‐1

DL (0‐2) DL (0‐2)b 1

0‐20‐30‐10‐1

DL (0‐2) DL (0‐2)b

2

0‐20‐20‐10‐10‐2

2

0‐20‐1

0

0‐10‐20‐20‐1

2

0‐3

DL (0‐1)

NR

DL (0‐1) DL (0‐1)

0‐1

DL (0‐1) DL (0‐1) DL (0‐1)

Visit Type

HAHAHA

ED

HAHAHAHA

EDED

HAHAHAHAHA

EDEDED

HA

EDEDEDED

HA

EDEDEDED

HA

ED

HAHA

ED

HAHAHAHAHAHA

‐25 ‐20 ‐15 ‐10 ‐5 0 5 10 15 20 25 

% Increase
3-35 

Studies examining the potential confounding effects of copollutants have reported that O3

effect estimates remained relatively robust upon the inclusion of PM and gaseous pollutants in 

two-pollutant models (U.S. 2013, Figure 6-20, Table 6-29). Additional studies that conducted 

copollutant analyses, but did not present quantitative results, also support these conclusions 

(Strickland et al., 2010; Tolbert et al., 2007; Medina-Ramon et al., 2006) (U.S. 2013, section 

6.2.7.5). 

In the last review, studies had not evaluated the concentration-response relationship 

between short-term O3 exposure and respiratory-related hospital admissions and emergency 

department visits. A preliminary examination of this relationship in studies that have become 

available since the last review found no evidence of a deviation from linearity when examining 

the association between short-term O3 exposure and asthma hospital admissions (U.S. EPA, 

2013, page 6-157, and Silverman and Ito, 2010). In addition, an examination of the 

concentration-response relationship for O3 exposure and pediatric asthma emergency department 

visits found no evidence of a threshold at O3 concentrations as low as 30 ppb (for daily 

maximum 8-hour concentrations) (Strickland et al., 2010). However, in both studies there is 

uncertainty in the shape of the concentration-response curve at the lower end of the distribution 

of O3 concentrations due to the low density of data in this range (U.S. 2013, page 6-157). 

Respiratory Mortality 

The controlled human exposure, epidemiologic, and toxicological studies discussed in 

section 6.2 of the ISA (U.S. EPA, 2013, section 6.2) provide strong evidence for respiratory 

morbidity effects, including ED visits and hospital admissions, in response to short-term O3

exposures. Moreover, evidence from experimental studies indicates multiple potential pathways 

of respiratory effects from short-term O3 exposures, which support the continuum of respiratory 

effects that could potentially result in respiratory-related mortality in adults (U.S. EPA, 2013, 

section 6.2.8). The 2006 O3 AQCD found inconsistent evidence for associations between shortterm O3 concentrations and respiratory mortality (U.S. EPA, 2006). Although some studies 

reported a strong positive association between O3 and respiratory mortality, additional studies 

reported small associations or no associations. New epidemiologic evidence for respiratory 

mortality is discussed in detail in section 6.2.8 of the ISA (U.S. EPA, 2013). The majority of 

recent multicity studies have reported positive associations between short-term O3 exposures and 

respiratory mortality, particularly during the summer months (U.S. EPA, 2013, Figure 6-36). 

Specifically, recent multicity studies from the U.S. (Zanobetti and Schwartz, 2008b), 

Europe (Samoli et al., 2009), Italy (Stafoggia et al., 2010), and Asia (Wong et al., 2010), as well 

as a multi-continent study (Katsouyanni et al., 2009), reported associations between short-term 

O3 concentrations and respiratory mortality (U.S. EPA, 2013, Figure 6-37, page 6-259). With 
3-36 

respect to respiratory mortality, summer-only analyses were consistently positive and most were 

statistically significant. In all-year analyses associations were positive, but smaller in magnitude. 

Of the studies evaluated, only the studies by Katsouyanni et al. (2009) and by Stafoggia 

et al. (2010) analyzed the potential for copollutant confounding of the O3-respiratory mortality 

relationship. Based on the results of these analyses, the ISA concluded that O3 respiratory 

mortality risk estimates appear to be moderately to substantially sensitive (e.g., increased or 

attenuated) to inclusion of PM10. However, in the APHENA study (Katsouyanni et al., 2009), the 

mostly every-6th-day sampling schedule for PM10 in the Canadian and U.S. datasets greatly 

reduced their sample size and limits the interpretation of these results (U.S. EPA, 2013, section 

6.2.8). 

In summary, recent epidemiologic studies support and reinforce the epidemiologic 

evidence for O3-associated respiratory hospital admissions and emergency department visits 

from the last review. In addition, the evidence for associations with respiratory mortality has 

been strengthened considerably since the last review, with the addition of several large multicity 

studies. The biological plausibility of the associations reported in these studies is supported by 

the experimental evidence for respiratory effects. 

3.1.2.2 Respiratory Effects – Long-term Exposures 

 To what extent does the currently available scientific evidence, including related 

uncertainties, strengthen or alter our understanding from the last review of 

respiratory effects attributable to long-term O3 exposures? 

As recognized in section 3.1.2.1, “the clearest evidence for health effects associated with 

exposure to O3 is provided by studies of respiratory effects” (U.S. EPA, 2013, section 1, p. 1-6). 

Collectively, there is a vast amount of evidence spanning several decades that supports a causal 

association between exposure to O3 and a continuum of respiratory effects (U.S. EPA, 2013, 

section 2.5). While the majority of this evidence is derived from studies investigating short-term 

exposures, evidence from animal toxicological studies and recent epidemiologic evidence 

indicate that long-term exposures (i.e., months to years) may also be detrimental to the 

respiratory system. Across this evidence, particularly the epidemiologic evidence, the exposures 

of focus vary, often involving repeated short concentrations extending over a long period, rather 

than a continuous long-term exposure period. 

In the 2006 O3 AQCD, evidence was examined for relationships between long-term O3

exposure and effects on respiratory health outcomes including declines in lung function, 

increases in inflammation, and development of asthma in children and adults. Animal toxicology 

data provided a clearer picture indicating that long-term O3 exposure may have lasting effects. 
3-37 

Chronic32 exposure studies in animals have reported biochemical and morphological changes 

suggestive of irreversible long-term O3 impacts on the lung. In contrast to supportive evidence 

from chronic animal studies, the epidemiologic studies on longer-term (annual) lung function 

declines, inflammation, and new asthma development remained inconclusive. 

Several epidemiologic studies collectively indicated that O3 exposure averaged over 

several summer months was associated with smaller increases in lung function growth in 

children. For longer averaging periods (annual), the analysis in the Children’s Health Study 

(CHS) reported by Gauderman et al. (2004) provided little evidence that such long-term 

exposure to ambient O3 was associated with significant deficits in the growth rate of lung 

function in children. Limited epidemiologic research examined the relationship between longterm O3 exposures and inflammation. Cross-sectional studies detected no associations between 

long-term O3 exposures and asthma prevalence, asthma-related symptoms or allergy to common 

aeroallergens in children. However, longitudinal studies provided evidence that long-term O3

exposure influences the risk of asthma development in children and adults. 

The currently available body of evidence supporting a relationship between long-term O3

exposures and adverse respiratory health effects that is likely to be causal is discussed in detail in 

the ISA (EPA 2013, section 7.2). New evidence reports interactions between genetic variants and 

long-term O3 exposure affect the occurrence of new-onset asthma in multi-community, U.S. 

cohort studies where protection by specific oxidant gene variants was restricted to children living 

in low O3 communities. A new line of evidence reports a positive concentration-response 

relationship between first asthma hospitalization and long-term O3 exposure. Related studies 

report coherent relationships between asthma severity and control, and respiratory symptoms 

among asthmatics and long-term O3 exposure. There is also limited evidence for an association 

between long-term exposure to ambient O3 concentrations and respiratory mortality. These 

studies are summarized briefly below for new-onset asthma and asthma prevalence, asthma 

hospital admissions and other morbidity effects, pulmonary structure and function, and 

respiratory mortality. 

Currently available scientific evidence of the adverse health effects attributable to longterm O3 exposures, even considering related uncertainties, is much stronger than the body of 

evidence available at the time of the 2008 review of the O3 standard. The 2006 O3 AQCD (U.S. 

EPA, 2006) concluded that epidemiologic studies provided no evidence of associations between 

long-term (annual) O3 exposures and asthma-related symptoms, asthma prevalence, or allergy to 

common allergens after controlling for covariates. It found limited evidence for a relationship 

 

32 Unless otherwise specified, the term “chronic” generally refers to an annual exposure duration for epidemiologic 

studies and a duration of greater than 10% of the lifespan of the animal in toxicological studies. 
3-38 

between long-term exposures to ambient O3 and deficits in the growth rate of lung-function in 

children, pulmonary inflammation and other endpoints. Episodic exposures were also known to 

cause more severe pulmonary morphological changes than continuous exposure. 

The evidence base available in this review includes additional epidemiologic studies 

using a variety of designs and analysis methods evaluating the relationship between long-term O3

exposures and measures of respiratory morbidity and mortality effects conducted by different 

research groups in different locations. The ISA (U.S. EPA, 2013, p. 7-33), in Table 7-2 presents 

selected key new longitudinal and cross-sectional studies of respiratory health effects and 

associated O3 concentrations. The positive results from various designs and locations support a 

relationship between long-term exposure to ambient O3 and respiratory health effects and 

mortality. 

 In this review, the evidence of effects associated with long-term exposures strengthens 

the relationship between O3 exposure and health effects defined as adverse by the ATS, a 

definition that has been used in previous reviews of the O3 standard. As discussed in more detail 

in section 3.1.3 below, the ATS (1985) defined adverse as “medically significant physiologic or 

pathologic changes generally evidenced by one or more of the following: (1) interference with 

the normal activity of the affected person or persons, (2) episodic respiratory illness, (3) 

incapacitating illness, (4) permanent respiratory injury, and/or (5) progressive respiratory 

dysfunction.” As discussed below, in this review there is now credible evidence of respiratory 

health effects associated with long-term O3 exposures that would fall in to each of these five 

categories that define adversity. 

From a policy perspective, the recent epidemiologic studies from the CHS of long-term 

O3 exposures that shed light on the interaction between genetic variability, O3 exposures, and 

health effects in children are important, not only because they help clarify previous findings, but 

also because the effects evaluated, such as new-onset asthma, are clearly adverse. The ISA (U.S. 

EPA, 2013, p. 7-12) notes that the collective evidence from CHS provides an important 

demonstration of gene-environment interactions. It further notes that in the complex 

gene-environment setting a modifying effect might not be reflected in an exposure main effect 

and that the simultaneous occurrence of main effect and interaction effect can occur. Moreover, 

the study of gene-environment interactions elucidates disease mechanisms in humans by using 

information on susceptibility genes to focus on the biological pathways that are most relevant to 

that disease. 

In the CHS cohort of children in 12 Southern California communities, long-term 

exposure to O3 concentrations was not associated with increased risk of developing asthma 

(McConnell et al., 2010); however, greater outdoor exercise was associated with development of 

asthma in children living in communities with higher ambient O3 concentrations (McConnell et 
3-39 

al., 2002). Recent CHS studies examined interactions among genetic variants, long-term O3

exposure, and new onset asthma in children. These prospective cohort studies are 

methodologically rigorous epidemiologic studies, and evidence indicates gene-O3 interactions. 

These studies have provided data supporting decreased risk of certain genetic variants on new 

onset asthma (e.g., HMOX-1, ARG) that is limited to children either in low (Islam et al., 2008) 

or high (Salam et al., 2009) O3 communities. Gene-environment interaction also was 

demonstrated with findings that greater outdoor exercise increased risk of asthma in GSTP1 

Ile/Ile children living in high O3 communities (Islam et al., 2009). Biological plausibility for 

these gene-O3 environment interactions is provided by evidence that these enzymes have 

antioxidant and/or anti-inflammatory activity and participate in well recognized modes of action 

in asthma pathogenesis. As O3 is a source of oxidants in the airways, oxidative stress serves as 

the link among O3 exposure, enzyme activity, and asthma. Cross-sectional studies by Akinbami 

et al. (2010) and Hwang et al. (2005) provide further evidence relating O3 exposures with asthma 

prevalence. 

Studies using a cross-sectional design provide support for a relationship between longterm O3 exposure and adverse health effects in asthmatics, including: bronchitic symptoms 

(related to TNF-308 genotype in asthmatic children) (Lee et al., 2009); asthma severity (Rage et 

al., 2009) and asthma control (Jacquemin et al., in press) in an adult cohort; respiratory-related 

school absences (related to CAT and MPO variant genes) (Wenten et al., 2009); asthma ED 

visits in adults (Meng et al., 2010); and, asthma hospital admissions in adults and children (Lin et 

al., 2008b; Meng et al., 2010; Moore et al., 2008). Several studies, shown in Table 7-3 (ISA, U.S. 

EPA, 2013, p. 7-35), provide results adjusted for potential confounders presenting results for 

both O3 and PM (in single and multipollutant models) as well as other pollutants where PM 

effects were not provided. As shown in this table, O3 associations were generally robust to 

adjustment by potential confounding by PM. 

Information from toxicological studies in nonhuman primates indicates that long term 

exposure to O3 during gestation or development can result in irreversible morphological changes 

in the lung, which in turn can influence the function of the respiratory tract. This nonhuman 

primate evidence of an O3-induced change in airway responsiveness supports the biologic 

plausibility of long term exposure to O3 contributing to effects of asthma in children. However, 

results from epidemiologic studies examining long-term O3 exposure and pulmonary function 

effects are inconclusive with some new studies relating effects at higher exposure levels. 

The ISA (U.S. EPA, 2013, p. 7-31) concludes that there is limited evidence for an 

association between long-term exposure to ambient O3 concentrations and respiratory mortality 

in adults (Jerrett et al., 2009). This effect was robust to the inclusion of PM2.5 and insensitive to a 

number of different model specifications. Moreover, there is evidence that long-term exposure to 
3-40 

O3 is associated with mortality among individuals that had previously experienced an emergency 

hospital admission due to COPD (Zanobetti and Schwartz, 2011). 

In conclusion, since the last review, the body of evidence about the effects of long-term 

O3 exposure has been considerably strengthened. The scientific evidence available for this 

review, including related uncertainties, provides an overall strong body of evidence of adverse 

health effects attributable to long-term O3 exposures. These include a coherent range of asthma 

morbidity effects such as new-onset asthma, asthma prevalence, symptoms, school absences, ED 

visits and hospital admissions. There is also new evidence of respiratory mortality associated 

with long-term O3 exposure. Further discussion of key studies is below. 

New-onset Asthma and Asthma Prevalence 

Asthma is a heterogeneous disease with a high degree of temporal variability. The on-set, 

progression, and symptoms can vary within an individual’s lifetime, and the course of asthma 

may vary markedly in young children, older children, adolescents, and adults. In the previous 

review, longitudinal cohort studies that examined associations between long-term O3 exposures 

and the onset of asthma in adults and children indicated a direct effect of long-term O3 exposures 

on asthma risk in adults (McDonnell et al., 1999, 15-year follow-up; Greer et al., 1993, 10-year 

follow-up) and effect modification by O3 in children (McConnell et al., 2002). Since that review, 

new evidence has become available about the association between long-term exposures to O3 and 

new-onset asthma that has increased our understanding of the gene-environment interaction and 

the mechanisms and biological pathways most relevant to assessing O3-related effects. 

In children, the relationship between long-term O3 exposure and new-onset asthma has 

been extensively studied in the CHS; a long-term study that was initiated in the early1990’s 

which has evaluated effects in several cohorts of children. The CHS was initially designed to 

examine whether long-term exposure to ambient pollution was related to chronic respiratory 

outcomes in children in 12 communities in southern California. In the CHS, new-onset asthma 

was classified as having no prior history of asthma at study entry with subsequent report of 

physician-diagnosed asthma at follow-up, with the date of onset assigned to be the midpoint of 

the interval between the interview date when asthma diagnosis was first reported and the 

previous interview date. The results of one study (McConnell et al., 2002) available in the 

previous review indicated that within high O3 communities, asthma risk was 3.3 times greater for 

children who played three or more outdoor sports as compared with children who played no 

sports. 

For this review, as discussed in section 7.2.1.1 of the ISA (U.S. EPA, 2013), recent 

studies from the CHS provide evidence for gene-environment interactions in effects on newonset asthma by indicating that the lower risks associated with specific genetic variants are found 

in children who live in lower O3 communities. These studies indicate that the risk for new-onset 
3-41 

asthma is related in part to genetic susceptibility, as well as behavioral factors and environmental 

exposure. The onset of a chronic disease, such as asthma, is partially the result of a sequence of 

biochemical reactions involving exposures to various environmental agents metabolized by 

enzymes related to a number of different genes. Oxidative stress has been proposed to underlie 

the mechanistic hypotheses related to O3 exposure. Genetic variants may impact disease risk 

directly, or modify disease risk by affecting internal dose of pollutants and other environmental 

agents and/or their reaction products, or by altering cellular and molecular modes of action. 

Understanding the relation between genetic polymorphisms and environmental exposure can 

help identify high-risk subgroups in the population and provide better insight into pathway 

mechanisms for these complex diseases. 

The CHS analyses (Islam et al., 2008; Islam et al. 2009; Salam et al., 2009) have found 

that asthma risk is related to interactions between O3 and variants in genes for enzymes such as 

heme-oxygenase (HO-1), arginases (ARG1 and 2), and glutathione S transferase P1 (GSTP1). 

Biological plausibility for these findings is provided by evidence that these enzymes have 

antioxidant and/or anti-inflammatory activity and participate in well-recognized modes of action 

in asthma pathogenesis. Further, several lines of evidence demonstrate that secondary oxidation 

products of O3 initiate the key modes of action that mediate downstream health effects (ISA, 

Section 5.3, U.S. EPA, 2013). For example, HO-1 responds rapidly to oxidants, has antiinflammatory and anti-oxidant effects, relaxes airway smooth muscle, and is induced in the 

airways during asthma. Gene-environment interactions are discussed in detail in Section 5.4.2.1 

in the ISA (U.S. EPA, 2013). 

Asthma Hospital Admissions 

In the 2006 AQCD, studies on O3-related hospital discharges and emergency department 

(ED) visits for asthma and respiratory disease mainly looked at short-term (daily) metrics. The 

short-term O3 studies presented in section 6.2.7.5 of the ISA (U.S. EPA, 2013) and discussed 

above in section 3.1.2.1 continue to indicate that there is evidence for increases in both hospital 

admissions and ED visits in children and adults related to all respiratory outcomes, including 

asthma, with stronger associations in the warm months. New studies, discussed in section 7.2.2 

of the ISA (U.S. EPA, 2013) also evaluated long-term O3 exposure metrics, providing a new line 

of evidence that suggests a positive exposure-response relationship between the first hospital 

admission for asthma and long-term O3 exposure, although the ISA cautions in attributing the 

associations in that study to long-term exposures since there is potential for short-term exposures 

to contribute to the observed associations. 

Evidence associating long-term O3 exposure to first asthma hospital admission in a 

positive concentration-response relationship is provided in a retrospective cohort study (Lin et 

al., 2008b). This study investigated the association between chronic exposure to O3 and 
3-42 

childhood asthma admissions by following a birth cohort of more than 1.2 million babies born in 

New York State (1995-1999) to first asthma admission or until 31 December 2000. Three annual 

indicators (all 8-hour maximum from 10:00 a.m. to 6:00 p.m.) were used to define chronic O3

exposure: (1) mean concentration during the follow-up period (41.06 ppb); (2) mean 

concentration during the O3 season (50.62 ppb); and (3) proportion of follow-up days with O3

levels >70 ppb. The effects of co-pollutants were controlled, and interaction terms were used to 

assess potential effect modifications. A positive association between chronic exposure to O3 and 

childhood asthma hospital admissions was observed, indicating that children exposed to high O3

levels over time are more likely to develop asthma severe enough to be admitted to the hospital. 

The various factors were examined and differences were found for younger children (1-2 years), 

poor neighborhoods, Medicaid/self-paid births, geographic region and others. As shown in the 

ISA, Figure 7-3 (U.S. EPA, 2013, p. 7-16), positive concentration-response relationships were 

observed. Asthma admissions were significantly associated with increased O3 levels for all 

chronic exposure indicators. 

In considering the relationship between long-term pollutant exposures and chronic 

disease heath endpoints, where chronic pathologies are found with acute expression of chronic 

disease, Künzli (2012) hypothesizes that if the associations of pollution with events are much 

larger in the long-term studies, it provides some indirect evidence that air pollution increases the 

pool of subjects with chronic disease, and that more acute events are to be expected to be seen 

for higher exposures. The results of Lin et al. (2008b) for first asthma hospital admission, 

presented in Figure 7-3 (U.S. EPA, 2013, p. 7-16), show effects estimates that are larger than 

those reported in a study of childhood asthma hospital admission in New York state (Silverman 

and Ito, 2010), discussed in section 3.1.2.1 and 3.1.2.2 above. The ISA (U.S. EPA, 2013, p. 7-16) 

notes that this provides some support for the hypothesis that O3 exposure may not only have 

triggered the events but also increased the pool of asthmatic children, but cautions in attributing 

the associations in Lin et al. (2008b) study to long-term exposures since there is potential for 

short-term exposures to contribute to the observed associations. 

Pulmonary structure and function 

In the 2006 O3 AQCD, few epidemiologic studies had investigated the effect of chronic 

O3 exposure on pulmonary function. The definitive 8-year follow-up analysis of the first cohort 

of the CHS (U.S. EPA, 2013, section 7.2.3.1) provided little evidence that long-term exposure to 

ambient O3 was associated with significant deficits in the growth rate of lung function in 

children. The strongest evidence was for medium-term effects of extended O3 exposures over 

several summer months on lung function (FEV1) in children, i.e., reduced lung function growth 

being associated with higher ambient O3 levels. Short-term O3 exposure studies presented in ISA 

(U.S. EPA, 2013, Section 6.2.1.2), and above in section 3.1.2.1, provide a cumulative body of 
3-43 

epidemiologic evidence that strongly supports associations between ambient O3 exposure and 

decrements in lung function among children. A later CHS study (Islam et al., 2007) included in 

this review (U.S. EPA, 2013, section 7.2.3.1) also reported no substantial differences in the 

effect of O3 on lung function. However, in a more recent CHS study, Breton et al. (2011) 

hypothesized that genetic variation in genes on the glutathione metabolic pathway may influence 

the association between ambient air pollutant exposures and lung function growth in children, 

and found that variation in the GSS locus was associated with differences in risk of children for 

lung function growth deficits associated ambient air pollutants, including O3. A recent study 

(Rojas-Martinez et al., 2007) of long-term exposure to O3, described in section 7.2.3.1 of the ISA 

(U.S. EPA, 2013, p. 7-19), observed a relationship with pulmonary function declines in schoolaged children where O3 and other pollutant levels were higher (90 ppb at high end of the range) 

than those in the CHS. Two studies of adult cohorts provide mixed results where long-term 

exposures were at the high end of the range. 

Long-term studies in animals allow for greater insight into the potential effects of 

prolonged exposure to O3 that may not be easily measured in humans, such as structural changes 

in the respiratory tract. Despite uncertainties, epidemiologic studies observing associations of O3

exposure with functional changes in humans can attain biological plausibility in conjunction with 

long-term toxicological studies, particularly O3-inhalation studies performed in non-human 

primates whose respiratory systems most closely resembles that of the human. An important 

series of studies, discussed in section 7.2.3.2 of the ISA (U.S. EPA, 2013), have used nonhuman 

primates to examine the effect of O3 alone, or in combination with an inhaled allergen, house 

dust mite antigen (HDMA), on morphology and lung function. These animals exhibit the 

hallmarks of allergic asthma defined for humans, including: a positive skin test for HDMA with 

elevated levels of IgE in serum and IgE-positive cells within the tracheobronchial airway walls; 

impaired airflow which is reversible by treatment with aerosolized albuterol; increased 

abundance of immune cells, especially eosinophils, in airway exudates and bronchial lavage; and 

development of nonspecific airway responsiveness (NHLBI, 2007). These studies and others 

have demonstrated changes in pulmonary function and airway morphology in adult and infant 

nonhuman primates repeatedly exposed to environmentally relevant concentrations of O3 (ISA, 

U.S. EPA, 2013, section 7.2.3.2). 

The initial observations in adult nonhuman primates have been expanded in a series of 

experiments using infant rhesus monkeys repeatedly exposed to 0.5 ppm O3 starting at 1 month 

of age (Plopper et al., 2007). The purpose of these studies was to determine if a cyclic regimen of 

O3 inhalation would amplify the allergic responses and structural remodeling associated with 

allergic sensitization and inhalation in the infant rhesus monkey. After several episodic 

exposures of infant monkeys to O3, they observed a significant increase in the baseline airway 
3-44 

resistance, which was accompanied by a small increase in airway responsiveness to inhaled 

histamine (Schelegle et al., 2003), although neither measurement was statistically different from 

filtered air control values. Exposure of animals to inhaled house dust mite antigen alone also 

produced small but not statistically significant changes in baseline airway resistance and airway 

responsiveness, whereas the combined exposure to both (O3 + antigen) produced statistically 

significant and greater than additive changes in both functional measurements. This nonhuman 

primate evidence of an O3-induced change in airway resistance and responsiveness provides 

biological plausibility of long-term exposure, or repeated short-term exposures, to O3

contributing to the effects of asthma in children. 

To understand which conducting airways and inflammatory mechanisms are involved in 

O3-induced airway hyperresponsiveness in the infant rhesus monkey, results of a follow-up study 

(Joad et al., 2006) suggest that effect of O3 on airway responsiveness occurs predominantly in 

the smaller bronchioles, where dosimetric models indicate the dose would be higher. 

The functional changes in the conducting airways were accompanied by a number of cellular and 

morphological changes, including a significant 4-fold increase in eosinophils. Thus, these studies 

demonstrate both functional and cellular changes in the lung of infant monkeys after cyclic 

exposure to 0.5 ppm O3, providing relevant information to understanding the potentially 

damaging effects of ambient O3 exposure on the respiratory tract of children. 

In addition, noteworthy structural changes in the respiratory tract development, during 

which conducting airways increase in diameter and length, have been observed in infant rhesus 

monkeys after cyclic exposure to O3 (Fanucchi et al., 2006). Observed changes included more 

proximal first alveolar outpocketing, decreases in the diameter and length of the terminal and 

respiratory bronchioles, increases in mucus-producing goblet cell mass, alterations in smooth 

muscle orientation in the respiratory bronchioles, epithelial nerve fiber distribution, and 

basement membrane zone morphometry. The latter effects are important because of their 

potential contribution to airway obstruction and airway hyperresponsiveness which are central 

features of asthma. A number of studies in both non-human primates and rodents demonstrate 

that O3 exposure can increase collagen synthesis and deposition, including fibrotic-like changes 

in the lung (ISA, U.S. EPA, 2013, section 7.2.3.2,). 

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 

and leading to apparent irreversible changes. Potentially, persistent inflammation and interstitial 

remodeling play an important role in the progression and development of chronic lung disease. 

Further discussion of the modes of action that lead to O3-induced morphological changes can be 

found in Section 5.3.7 of the ISA (U.S. EPA, 2013). Discussion of mechanisms involved in 

lifestage susceptibility and developmental effects can be found in Section 5.4.2.4 of the ISA 
3-45 

(U.S. EPA, 2013). The findings reported in chronic animal studies offer 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 (see Section 7.2.3.1). 

Respiratory Mortality 

 A limited number of epidemiologic studies have assessed the relationship between longterm exposure to O3 and mortality in adults. The 2006 O3 AQCD concluded that an insufficient 

amount of evidence existed “to suggest a causal relationship between chronic O3 exposure and 

increased risk for mortality in humans” (U.S. EPA, 2006). Though total and cardio-pulmonary 

mortality were considered in these studies, respiratory mortality was not specifically considered. 

In the most recent follow-up analysis of the ACS cohort (Jerrett et al., 2009), cardiopulmonary 

deaths were separately subdivided into respiratory and cardiovascular deaths, rather than 

combined as in the Pope et al. (2002) work. Increased O3 exposure was associated with the risk 

of death from respiratory causes, and this effect was robust to the inclusion of PM2.5. The 

association between increased O3 concentrations and increased risk of death from respiratory 

causes was insensitive to the use of different models and to adjustment for several ecologic 

variables considered individually. Additionally, a recent multi-city time series study (Zanobetti 

and Schwartz, 2011), which followed (from 1985 to 2006) four cohorts of Medicare enrollees 

with chronic conditions that might predispose to O3-related effects, observed an association 

between long-term (warm season) exposure to O3 and elevated risk of mortality in the cohort that 

had previously experienced an emergency hospital admission due to COPD. A key limitation of 

this study is the inability to control for PM2.5, because data were not available in these cities until 

1999. 

3.1.2.3 Total Mortality – Short-term Exposures 

 To what extent does the currently available scientific evidence, including related 

uncertainties, strengthen or alter our understanding from the last review of 

mortality attributable to short-term O3 exposures? 

The 2006 O3 AQCD concluded that the overall body of evidence was highly suggestive 

that short-term exposure to O3 directly or indirectly contributes to nonaccidental and 

cardiopulmonary-related mortality in adults, but additional research was needed to more fully 

establish underlying mechanisms by which such effects occur (U.S. EPA, 2006; U.S. EPA, 2013, 

p. 2-18). In building on the 2006 evidence, the ISA states the following (U.S. EPA, 2013, p. 6-

261). 

The evaluation of new multicity studies that examined the association between 

short-term O3 exposures and mortality found evidence that supports the 

conclusions of the 2006 AQCD. These new studies reported consistent positive 

associations between short-term O3 exposure and all-cause (non-accidental) 
3-46 

mortality, with associations persisting or increasing in magnitude during the 

warm season, and provide additional support for associations between O3

exposure and cardiovascular and respiratory mortality 

The 2006 O3 AQCD reviewed a large number of time-series studies of associations 

between short-term O3 exposures and total mortality including single- and multicity studies, and 

meta-analyses. In the large U.S. multicity studies that examined all-year data, effect estimates 

corresponding to single-day lags ranged from a 0.5-1% increase in all-cause (nonaccidental) total 

mortality per a 20 ppb (24-hour), 30 ppb (8-hour maximum), or 40 ppb (1-hour maximum) 

increase in ambient O3 (U.S. EPA, 2013, section 6.6.2). Available studies reported some 

evidence for heterogeneity in O3 mortality risk estimates across cities and across studies. Studies 

that conducted seasonal analyses reported larger O3 mortality risk estimates during the warm 

season. Overall, the 2006 O3 AQCD identified robust associations between various measures of 

daily ambient O3 concentrations and all-cause mortality, which could not be readily explained by 

confounding due to time, weather, or copollutants. With regard to cause-specific mortality, 

consistent positive associations were reported between short-term O3 exposure and 

cardiovascular mortality, with less consistent evidence for associations with respiratory 

mortality. The majority of the evidence for associations between O3 and cause-specific mortality 

were from single-city studies, which had small daily mortality counts and subsequently limited 

statistical power to detect associations. The 2006 O3 AQCD concluded that “the overall body of 

evidence is highly suggestive that O3 directly or indirectly contributes to non-accidental and 

cardiopulmonary-related mortality” (U.S. EPA, 2013, section 6.6.1). 

Recent studies have strengthened the body of evidence that supports the association 

between short-term O3 concentrations and mortality in adults. This evidence includes a number 

of studies reporting associations with non-accidental as well as cause-specific mortality. Multicontinent and multicity studies have consistently reported positive and statistically significant 

associations between short-term O3 concentrations and all-cause mortality, with evidence for 

larger mortality risk estimates during the warm or summer months (Figure 3-3 below, reprinted 

from the ISA) (U.S. EPA, 2013, Figure 6-27; Table 6-42). Similarly, evaluations of causespecific mortality have reported consistently positive associations with O3, particularly in 

analyses restricted to the warm season (U.S. EPA, 2013, Figure 6-37; Table 6-53).33 

 

33 Respiratory mortality is discussed in more detail above. 
3-47 

Figure 3-3. Summary of mortality risk estimates for short-term O3 and all-cause (nonaccidental) mortalit34 Reprinted from the ISA (U.S. EPA, 2013, Figure 6-27). 

Study

Gryparis et al. (2004;57276)

Bell et al. (2007; 93256)

Schwartz(2005;57333)

Bell and Dominici(2008;193828)

Bell et al. (2004; 94417)a

Levy et al. (2005; 74347)a

Katsouyanni et al. (2009; 199899)

Bell et al. (2005; 74345)a

Ito et al. (2005; 74346)a

Wong et al. (2010; 732535)

Katsouyanni et al. (2009; 199899)

Cakmak et al. (2011, 699135)

Katsouyanni et al. (2009; 199899)

Katsouyanni et al. (2009; 199899)b

Samoli et al. (2009; 195855)

Bell et al. (2004; 94417)a

Schwartz(2005;57333)

Zanobetti and Schwartz(2008;195755)

Zanobetti and Schwartz(2008;101596)

Franklin and Schwartz(2008; 156448)

Gryparis et al. (2004;57276)

Medina‐Ramon and Schwartz(2008)

Katsouyanni et al. (2009; 199899)

Bell et al. (2005; 74345)a

Katsouyanni et al. (2009; 199899)

Katsouyanni et al. (2009; 199899)b

Levy et al. (2005; 74347)a

Ito et al. (2005; 74346)a

Katsouyanni et al. (2009; 199899)

Stafoggia et al. (2010; 625034)

Location

APHEA2 (23 cities) 98 U.S. communities

14 U.S. cities

98 U.S. communities

95 U.S. communities

U.S. and Non‐U.S.

APHENA‐Europe U.S. and Non‐U.S.

U.S. and Non‐U.S.

PAPA (4 cities) APHENA‐U.S.

7 Chilean cities

APHENA‐Canada

APHENA‐Canada

21 European cities 95 U.S. communities

14 U.S. cities

48 U.S. cities

48 U.S. cities

18 U.S. communities

APHEA2 (21 cities) 48 U.S. cities

APHENA‐Europe U.S. and Non‐U.S.

APHENA‐Canada

APHENA‐Canada

U.S. and Non‐U.S.U.S. and Non‐U.S.

APHENA‐U.S.

10 Italian cities

Lag

0‐10‐1

0

0‐60‐6

DL(0‐2)

0‐1

DL(0‐2) DL(0‐6) DL(0‐2) DL(0‐2)

0‐10‐6

00

0‐3

0

0‐10‐2

DL(0‐2)DL(0‐2) DL(0‐2)DL(0‐2) DL(0‐5) 

‐1 1 3 5 7 

% Increase
3-48 

In assessing the evidence for O3-related mortality, the 2006 AQCD also noted that 

multiple uncertainties remained regarding the relationship between short-term O3 concentrations 

and mortality, including the extent of residual confounding by co-pollutants; characterization of 

the factors that modify the O3-mortality association; the appropriate lag structure for identifying 

O3-mortality effects; and the shape of the O3-mortality concentration-response function and 

whether a threshold exists. Many of the studies, published since the last review, have attempted 

to address one or more of these uncertainties. The ISA (U.S. EPA, 2013, Section 6.6.2) discusses 

the extent to which recent studies have evaluated these uncertainties in the relationship between 

O3 and mortality. 

In particular, recent studies have evaluated different statistical approaches to examine the 

shape of the O3-mortality concentration-response relationship and to evaluate whether a 

threshold exists for O3-related mortality. In an analysis of the NMMAPS data, Bell et al. (2006) 

evaluated the potential for a threshold in the O3-mortality relationship. The authors reported 

positive and statistically significant associations with mortality in a variety of subset analyses, 

including analyses restricted to days with 24-hour area-wide average O3 concentrations below 

60, 55, 50, 45, 40, 35, and 30 ppb. In these restricted analyses O3 effect estimates were of similar 

magnitude, were statistically significant, and had similar statistical precision. In analyses 

restricted to days with 24-hour average O3 concentrations below 25 ppb, the O3 effect estimate 

was similar in magnitude to the effect estimates resulting from analyses with the higher cutoffs, 

but had somewhat lower statistical precision, with the estimate approaching statistical 

significance (i.e., based on observation of Figure 2 in Bell et al., 2006). In analyses restricted to 

days with lower 24-hour average O3 concentrations, effect estimates were not statistically 

significant (i.e., based on observation of Figure 2 in Bell et al., 2006). 

Bell et al. (2006) also evaluated the shape of the concentration-response relationship 

between O3 and mortality. Although the results of this analysis suggested the lack of threshold in 

the O3-mortality relationship, the ISA noted that it is difficult to interpret such a curve because: 

(1) there is uncertainty around the shape of the concentration-response curve at 24-hour average 

O3 concentrations generally below 20 ppb and (2) the concentration-response curve does not take 

into consideration the heterogeneity in O3-mortality risk estimates across cities (U.S. EPA, 2013, 

section 6.6.2.3). 

Several additional studies have used the NMMAPS dataset to evaluate the concentrationresponse relationship between short-term O3 concentrations and mortality. For example, using 

the same data as Bell et al. (2006), Smith et al. (2009) conducted a subset analysis, but instead of 

restricting the analysis to days with O3 concentrations below a cutoff the authors only included 

days above a defined cutoff. The results of this analysis were consistent with those reported by 

Bell et al. (2006). Specifically, the authors reported consistent positive associations for all cutoff 
3-49 

concentrations up to concentrations where the total number of days available were so limited that 

the variability around the central estimate was increased (U.S. EPA, 2013, section 6.6.2.3). In 

addition, using NMMAPS data for 1987-1994 for Chicago, Pittsburgh, and El Paso, Xia and 

Tong (2006) reported evidence for a threshold around a 24-hour average O3 concentration of 

25 ppb, though the threshold values estimated in the analysis were sometimes in the range of 

where data density was low (U.S. EPA, 2013, section 6.6.2.3). Stylianou and Nicolich (2009) 

examined the existence of thresholds following an approach similar to Xia and Tong (2006) 

using data from NMMAPS for nine major U.S. cities (i.e., Baltimore, Chicago, Dallas/Fort 

Worth, Los Angeles, Miami, New York, Philadelphia, Pittsburgh, and Seattle) for the years 

1987-2000. The authors reported that the estimated O3-mortality risks varied across the nine 

cities, with the models exhibiting apparent thresholds in the 10-45 ppb range for O3 (24-hour 

average). Additional studies in Europe, Canada, and Asia did not report evidence for a threshold 

(Katsouyanni et al., 2009). 

3.1.2.4 Cardiovascular effects – Short-term Exposure 

 To what extent does the currently available scientific evidence, including related 

uncertainties, strengthen or alter our understanding from the last review of 

cardiovascular effects attributable to short-term O3 exposures? 

A relatively small number of studies have examined the potential effect of short-term O3

exposure on the cardiovascular system. The 2006 O3 AQCD (U.S. EPA, 2006, p. 8-77) 

concluded that “O3 directly and/or indirectly contributes to cardiovascular-related morbidity” but 

added that the body of evidence was limited. This conclusion was based on a controlled human 

exposure study that included hypertensive adult males; a few epidemiologic studies of 

physiologic effects, heart rate variability, arrhythmias, myocardial infarctions, and hospital 

admissions; and toxicological studies of heart rate, heart rhythm, and blood pressure. 

More recently, the body of scientific evidence available that has examined the effect of 

O3 on the cardiovascular system has expanded. There is an emerging body of animal 

toxicological evidence demonstrating that short-term exposure to O3 can lead to autonomic 

nervous system alterations (in heart rate and/or heart rate variability) and suggesting that 

proinflammatory signals may mediate cardiovascular effects. Interactions of O3 with respiratory 

tract components result in secondary oxidation product formation and subsequent production of 

inflammatory mediators, which have the potential to penetrate the epithelial barrier and to initiate 

toxic effects systemically. In addition, animal toxicological studies of long-term exposure to O3

provide evidence of enhanced atherosclerosis and ischemia/reperfusion (I/R) injury, 

corresponding with development of a systemic oxidative, proinflammatory environment. Recent 

experimental and epidemiologic studies have investigated O3-related cardiovascular events and 
3-50 

are summarized in Section 6.3 of the ISA (U.S. EPA, 2013, Section 6.3). Overall, the ISA 

summarized the evidence in this review as follows (U.S. EPA, 2013, p. 6-211). 

In conclusion, animal toxicological studies demonstrate O3-induced 

cardiovascular effects, and support the strong body of epidemiologic evidence 

indicating O3-induced cardiovascular mortality. Animal toxicological and 

controlled human exposure studies provide evidence for biologically plausible 

mechanisms underlying these O3-induced cardiovascular effects. However, a lack 

of coherence with epidemiologic studies of cardiovascular morbidity remains an 

important uncertainty.

Animal toxicological studies support that short-term O3 exposure can lead to 

cardiovascular morbidity. Animal studies provide evidence for both increased and decreased 

heart rate (HR), however it is uncertain if O3-induced reductions in HR are relevant to humans. 

Animal studies also provide evidence for increased heart rate variability (HRV), arrhythmias, 

vascular disease and injury following short-term O3 exposure. In addition, a series of studies 

highlight the role of genetic variability and age in the induction of effects and attenuation of 

responses to O3 exposure. 

Biologically plausible mechanisms have been described for the cardiovascular effects 

observed in animal exposure studies (U.S. EPA, 2013, Section 5.3.8). Evidence that 

parasympathetic pathways may underlie cardiac effects is described in more detail in Section 

5.3.2 of the ISA (U.S. EPA, 2013). Recent studies suggest that O3 exposure may disrupt the 

endothelin system that constricts blood vessels and increase blood pressure, which can result in 

an increase in HR, HRV; and disrupt the NO ̇system and the production of atrial natriuretic 

factor (ANF), vasodilators that reduce blood pressure. Additionally, O3 may increase oxidative 

stress and vascular inflammation promoting the progression of atherosclerosis and leading to 

increased susceptibility to I/R injury. As O3 reacts quickly with the ELF and does not translocate 

to the heart and large vessels, studies suggest that the cardiovascular effects exhibited could be 

caused by secondary oxidation products resulting from O3 exposure. However, direct evidence of 

translocation of O3 reaction products to the cardiovascular system has not been demonstrated in 

vivo. Alternatively, extrapulmonary release of diffusible mediators (such as cytokines or 

endothelins) may initiate or propagate inflammatory responses throughout the body leading to 

the cardiovascular effects reported in toxicology studies. Ozone reacts within the lung to induce 

pulmonary inflammation and the influx and activation of inflammatory cells, resulting in a 

cascading proinflammatory state, and may lead to the extrapulmonary release of diffusible 

mediators that could result in cardiovascular injury. 

Controlled human exposures studies discussed in previous AQCDs have not 

demonstrated any consistent extrapulmonary effects. In this review, evidence from controlled 

human exposure studies suggests cardiovascular effects in response to short-term O3 exposure 
3-51 

(see ISA, U.S. EPA, 2013, Section 6.3.1) and provides some coherence with evidence from 

animal toxicology studies. Controlled human exposure studies also support the animal 

toxicological studies by demonstrating O3-induced effects on blood biomarkers of systemic 

inflammation and oxidative stress, as well as changes in biomarkers that can indicate a 

prothrombogenic response to O3. Increases and decreases in high frequency HRV have been 

reported following relatively low (120 ppb during rest) and high (300 ppb with exercise) O3

exposures, respectively. These changes in cardiac function observed in animal and human studies 

provide preliminary evidence for O3-induced modulation of the autonomic nervous system 

through the activation of neural reflexes in the lung (see ISA, U.S. EPA, 2013, Section 5.3.2). 

Overall, the ISA concludes that the available body of epidemiologic evidence examining 

the relationship between short-term exposures to O3 concentrations and cardiovascular morbidity 

is inconsistent (U.S. EPA, 2013, Section 6.3.2.9). Across studies, different definitions, (i.e., ICD9 diagnostic codes) were used for both all-cause and cause-specific cardiovascular morbidity 

(ISA, U.S. EPA, 2013, see Tables 6-35 to 6-39), which may contribute to inconsistency in 

results. However, within diagnostic categories, no consistent pattern of association was found 

with O3. Generally, the epidemiologic studies used nearest air monitors to assess O3

concentrations, with a few exceptions that used modeling or personal exposure monitors. 

The inconsistencies in the associations observed between short-term O3 and cardiovascular 

disease (CVD) morbidities are unlikely to be explained by the different exposure assignment 

methods used (see Section 4.6, ISA, U.S. EPA, 2013). The wide variety of biomarkers 

considered and the lack of consistency among definitions used for specific cardiovascular disease 

endpoints (e.g., arrhythmias, HRV) make comparisons across studies difficult. 

Despite the inconsistent evidence for an association between O3 concentration and CVD 

morbidity, mortality studies indicate a consistent positive association between short-term O3

exposure and cardiovascular mortality in multicity studies and in a multicontinent study. When 

examining mortality due to cardiovascular disease, epidemiologic studies consistently observe 

positive associations with short-term exposure to O3. Additionally, there is some evidence for an 

association between long-term exposure to O3 and mortality, although the association between 

long-term ambient O3 concentrations and cardiovascular mortality can be confounded by other 

pollutants as evident by a study of cardiovascular mortality that reported no association after 

adjustment for PM2.5 concentrations. The ISA (U.S. EPA, 2013, section 6.3.4) states that taken 

together, the overall body of evidence across the animal and human studies is sufficient to 

conclude that there is likely to be a causal reationship between relevant short-term exposures to 

O3 and cardiovascular system effects. 
3-52 

3.1.3 Adversity of Effects 

In this section we address the following question: 

 To what extent does the currently available scientific evidence expand our 

understanding of the adversity of O3-related health effects? 

In making judgments as to when various O3-related effects become regarded as adverse 

to the health of individuals, in previous NAAQS reviews staff has relied upon the guidelines 

published by the ATS and the advice of CASAC. In 2000, the ATS published an official 

statement on “What Constitutes an Adverse Health Effect of Air Pollution?” (ATS, 2000), which 

updated and built upon its earlier guidance (ATS, 1985). The earlier guidance defined adverse 

respiratory health effects as “medically significant physiologic changes generally evidenced by 

one or more of the following: (1) interference with the normal activity of the affected person or 

persons, (2) episodic respiratory illness, (3) incapacitating illness, (4) permanent respiratory 

injury, and/or (5) progressive respiratory dysfunction”, while recognizing that perceptions of 

“medical significance” and “normal activity” may differ among physicians, lung physiologists 

and experimental subjects (ATS, 1985). The 2000 ATS guidance builds upon and expands the 

1985 definition of adversity in several ways. The guidance concludes that transient, reversible 

loss of lung function in combination with respiratory symptoms should be considered adverse. 

There is also a more specific consideration of population risk (ATS, 2000). Exposure to air 

pollution that increases the risk of an adverse effect to the entire population is adverse, even 

though it may not increase the risk of any individual to an unacceptable level. For example, a 

population of asthmatics could have a distribution of lung function such that no individual has a 

level associated with clinical impairment. Exposure to air pollution could shift the distribution to 

lower levels that still do not bring any 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 to 

further environmental insult (U.S. EPA, 2013, p. lxxi; and 75 FR at 35526/2, June 22, 2010). 

The ATS also concluded that elevations of biomarkers such as cell types, cytokines and 

reactive oxygen species may signal risk for ongoing injury and more serious effects or may 

simply represent transient responses, illustrating the lack of clear boundaries that separate 

adverse from nonadverse events. More subtle health outcomes also may be connected 

mechanistically to health effects that are clearly adverse, so that small changes in physiological 

measures may not appear clearly adverse when considered alone, but may be part of a coherent 

and biologically plausible chain of related health outcomes that include responses that are clearly 

adverse, such as mortality (section 3.1.2.1, above). 
3-53 

In this review, the new evidence provides further support for relationships between O3

exposures and a spectrum of health effects, including effects that meet the ATS criteria for being 

adverse (ATS, 1985 and 2000). The ISA judgment that there is a causal relationship between 

short-term O3 exposure and a full range of respiratory effects, including respiratory morbidity 

(e.g., lung function decrements, respiratory symptoms, inflammation, hospital admissions, and 

emergency department visits) and mortality, provides support for concluding that short-term O3

exposure is associated with adverse effects (U.S. EPA, 2013, section 2.5.2). Overall, including 

new evidence of cardiovascular system effects, the evidence supporting an association between 

short-term O3 exposures and total (non-accidental, cardiopulmonary) respiratory mortality is 

stronger in this review (U.S. EPA, 2013, section 2.5.2). And the judgment of likely causal 

associations between long-term measures of O3 exposure and respiratory effects such as newonset asthma, prevalence of asthma, asthma symptoms and control, and asthma hospital 

admissions provides support for concluding that long-term O3 exposure is associated with 

adverse effects ranging from episodic respiratory illness to permanent respiratory injury or 

progressive respiratory decline (U.S. EPA, 2013, section 7.2.8). 

This review provides additional evidence of O3-attributable effects that are clearly 

adverse, including premature mortality. Application of the ATS guidelines to the least serious 

category of effects related to ambient O3 exposures, which are also the most numerous and 

therefore are also potentially important from a public health perspective, involves judgments 

about which medical experts on CASAC panels and public commenters have in the past 

expressed diverse views. To help frame such judgments, EPA staff defined gradations of 

individual functional responses (e.g., decrements in FEV1 and airway responsiveness) and 

symptomatic responses (e.g., cough, chest pain, wheeze), together with judgments as to the 

potential impact on individuals experiencing varying degrees of severity of these responses. 

These gradations were used in the 1997 O3 NAAQS review and slightly revised in the 2008 

review (U.S. EPA, 1996, p. 59; 2007, p. 3-72; 72 FR 37849, July 11, 2007). These gradations 

and impacts are summarized in Tables 3-2 and 3-3 in the 2007 O3 Staff Paper (U.S. EPA, 2007, 

pp. 3-74 to 3-75). 

For active healthy people, including children, moderate levels of functional responses 

(e.g., FEV1 decrements of ≥ 10% but < 20%, lasting 4 to 24 hours) and/or moderate symptomatic 

responses (e.g., frequent spontaneous cough, marked discomfort on exercise or deep breath, 

lasting 4 to 24 hours) would likely interfere with normal activity for relatively few sensitive 

individuals (U.S. EPA, 2007, p. 3-72; 72 FR 37849, July 11, 2007); whereas large functional 

responses (e.g., FEV1 decrements ≥ 20%, lasting longer than 24 hours) and/or severe 

symptomatic responses (e.g., persistent uncontrollable cough, severe discomfort on exercise or 

deep breath, lasting longer than 24 hours) would likely interfere with normal activities for many 
3-54 

sensitive individuals (U.S. EPA, 2007, p. 3-72; 72 FR 37849, July 11, 2007) and therefore would 

be considered adverse under ATS guidelines. For the purpose of estimating potentially adverse 

lung function decrements in active healthy people in the 2008 O3 NAAQS review, the CASAC 

panel for that review indicated that a focus on the mid to upper end of the range of moderate 

levels of functional responses is most appropriate (e.g., FEV1 decrements ≥ 15% but < 20%) 

(Henderson, 2006; U.S. EPA, 2007, p. 3-76). In this review, CASAC concurred that the 

“[e]stimation of FEV1 decrements of ≥15% is appropriate as a scientifically relevant surrogate 

for adverse health outcomes in active healthy adults” (Frey, 2014, p. 3). However, for children 

and adults with lung disease, even moderate functional (e.g., FEV1 decrements ≥ 10% but < 

20%, lasting up to 24 hours) or symptomatic responses (e.g., frequent spontaneous cough, 

marked discomfort on exercise or with deep breath, wheeze accompanied by shortness of breath, 

lasting up to 24 hours) would likely interfere with normal activity for many individuals, and 

would likely result in additional and more frequent use of medication (U.S. EPA, 2007, p.3-72; 

72 FR 37849, July 11, 2007). For people with lung disease, large functional responses (e.g., 

FEV1 decrements ≥ 20%, lasting longer than 24 hours) and/or severe symptomatic responses 

(e.g., persistent uncontrollable cough, severe discomfort on exercise or deep breath, persistent 

wheeze accompanied by shortness of breath, lasting longer than 24 hours) would likely interfere 

with normal activity for most individuals and would increase the likelihood that these individuals 

would seek medical treatment (U.S. EPA, 2007, p.3-72; 72 FR 37849, July 11, 2007). In the last 

O3 NAAQS review, for the purpose of estimating potentially adverse lung function decrements 

in people with lung disease the CASAC panel indicated that a focus on the lower end of the 

range of moderate levels of functional responses is most appropriate (e.g., FEV1 decrements 

≥10%) (Henderson, 2006; U.S. EPA, 2007, p. 3-76). In addition, in the reconsideration of the 

2008 final decision, CASAC stated that “[a] 10% decrement in FEV1 can lead to respiratory 

symptoms, especially in individuals with pre-existing pulmonary or cardiac disease. For 

example, people with chronic obstructive pulmonary disease have decreased ventilatory reserve 

(i.e., decreased baseline FEV1) such that a ≥10% decrement could lead to moderate to severe 

respiratory symptoms” (Samet, 2011) (section 3.1.2.1, above). In this review, CASAC concurred 

that “[a]n FEV1 decrement of ≥10% is a scientifically relevant surrogate for adverse health outcomes 

for people with asthma and lung disease” (Frey, 2014, p. 3). 

In judging the extent to which these impacts represent effects that should be regarded as 

adverse to the health status of individuals, in previous NAAQS reviews we also considered 

whether effects were experienced repeatedly during the course of a year or only on a single 

occasion (Staff Paper, U.S. EPA, 2007). Although some experts would judge single occurrences 

of moderate responses to be a ‘‘nuisance,’’ especially for healthy individuals, a more general 

consensus view of the adversity of such moderate responses emerges as the frequency of 
3-55 

occurrence increases. Thus it has been judged that repeated occurrences of moderate responses, 

even in otherwise healthy individuals, may be considered to be adverse since they could well set 

the stage for more serious illness (61 FR 65723). The CASAC panel in the 1997 NAAQS review 

expressed a consensus view that these “criteria for the determination of an adverse physiological 

response were reasonable” (Wolff, 1995). In the review completed in 2008, estimates of repeated 

occurrences continued to be an important public health policy factor in judging the adversity of 

moderate lung function decrements in healthy and asthmatic people (72 FR 37850, July 11, 

2007). 

Evidence new to this review indicates that 6.6-hour exposures to 60 ppb O3 during 

moderate exertion can result in pulmonary inflammation in healthy adults. As discussed in 

section 3.1.2 above, the initiation of inflammation can be considered as evidence that injury has 

occurred. Inflammation induced by a single O3 exposure can resolve entirely, but continued 

acute inflammation can evolve into a chronic inflammatory state (ISA, U.S. EPA, 2013, p. 6-76), 

which is clearly adverse. Therefore, like moderate lung function decrements, whether 

inflammation is experienced repeatedly during the course of a year or only on a single occasion 

is judged by staff to be reasonable criteria for determining adverse inflammatory effects 

attributable to O3 exposures at 60 ppb. 

Responses measured in controlled human exposure studies indicate that the range of 

effects elicited 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 (EPA, 2013, section 6.2.1.1). Some young, 

healthy adults exposed to O3 concentrations ≥ 60 ppb, while engaged in 6.6 hours of intermittent 

moderate exertion, develop reversible, transient decrements in lung function, symptoms of 

breathing discomfort, and inflammation if minute ventilation or duration of exposure is increased 

sufficiently (EPA, 2013, section 6.2.1.1). Among healthy subjects there is considerable 

interindividual variability in the magnitude of the FEV1 responses, but averaged across studies at 

60 ppb (EPA, 2013, pp. 6-17 to 6-18), 10% of healthy subjects had >10% FEV1 decrements. 

Moreover, consistent with the findings of the ISA (EPA, 2013, section 6.2.1.1), CASAC 

concluded that “[a]sthmatic subjects appear to be at least as sensitive, if not more sensitive, than 

non-asthmatic subjects in manifesting ozone-induced pulmonary function decrements” (Frey, 2014, 

p. 4). The combination of lung function decrements and respiratory symptoms, which has been 

considered adverse in previous reviews, has been demonstrated in healthy adults following 

prolonged (6.6 hour) exposures, while at intermittent moderate exertion, to 70 ppb. For these 

types of effects, information from controlled human exposure studies, which provides an 

indication of the magnitude and thus adversity of effects at different O3 concentrations, 
3-56 

combined with estimates of occurrences in the population from the HREA, provide information 

about their importance from a policy perspective. 

3.1.4 Ozone Concentrations Associated With Health Effects 

In evaluating O3 exposure concentrations reported to result in health effects, within the 

context of the adequacy of the current standard, we first consider the following specific question: 

 To what extent does the currently available scientific evidence indicate morbidity 

and/or mortality attributable to exposures to O3 concentrations lower than 

previously reported or that would meet the current standard?

In addressing this question, we characterize the extent to which O3-attributable effects have been 

reported over the ranges of O3 exposure concentrations evaluated in controlled human exposure 

studies and over the distributions of ambient O3 concentrations in locations where epidemiologic 

studies have been conducted. 

3.1.4.1 Concentrations in Controlled Human Exposure Studies and in Epidemiologic 

Panel Studies 

In considering what the currently available evidence indicates with regard to effects 

associated with exposure concentrations lower than those identified in the last review, or that 

could meet the current standard, we first consider the evidence from controlled human exposure 

studies and epidemiologic panel studies. This evidence is assessed in section 6.2 of the ISA and 

is summarized in section 3.1.2 above. Controlled human exposure studies have generally been 

conducted with young, healthy adults, and have evaluated exposure durations less than 8 hours. 

Epidemiologic panel studies have evaluated a wider range of study populations, including 

children, and have generally evaluated associations with O3 concentrations averaged over several 

hours (U.S. EPA, 2013, section 6.2.1.2).35 

As summarized above (section 3.1.2.1), and as discussed in detail in the ISA (U.S. EPA, 

2013, section 6.2), a large number of controlled human exposure studies have reported lung 

function decrements, respiratory symptoms, airway inflammation, airway hyperresponsiveness, 

and/or impaired lung host defense in young, healthy adults engaged in moderate, intermittent 

exertion, following 6.6-hour O3 exposures. These studies have consistently reported such effects 

following exposures to O3 concentrations of 80 ppb or greater. Available studies have also 

evaluated some of these effects (i.e., lung function decrements, respiratory symptoms, airway 

inflammation) following exposures to O3 concentrations below 75 ppb. Table 3-1 highlights the 

 

35 In this section we focus on panel studies that used on-site monitoring, and that are highlighted in the ISA for the 

extent to which monitored ambient O3 concentrations reflect exposure concentrations in their study populations 

(U.S. EPA, 2013, section 6.2.1.2). 
3-57 

group mean results of individual controlled human exposure studies that have evaluated 

exposures of healthy adults to O3 concentrations below 75 ppb. The studies included in Table 3-1 

indicate lung function decrements, airway inflammation, and respiratory symptoms in healthy 

adults following exposures to O3 concentrations below 75 ppb. 
3-58 

Table 3-1. Group mean results of controlled human exposure studies that have evaluated 

exposures to ozone concentrations below 75 ppb in young, healthy adults. 

Endpoint O3 Exposure 

Concentration Study 

Statistically 

Significant O3-

Induced 

Effect36

FEV1 decrements 

70 ppb Schelegle et al., 200937 yes 

60 ppb 

Kim et al., 2011 yes 

Schelegle et al., 200938 no 

Adams, 2006 yes39

Adams, 2002 no 

40 ppb Adams, 2006 no 

Adams, 2002 no 

Respiratory 

Symptoms 

70 ppb Schelegle et al., 2009 yes 

60 ppb 

Kim et al., 2011 no 

Schelegle et al., 2009 no 

Adams, 2006 no40

40 ppb Adams, 2006 no 

Adams, 2002 no 

Airway 

Inflammation 

(neutrophil influx) 

60 ppb Kim et al., 2011 yes 

 In further evaluating O3-induced FEV1 decrements following exposures to O3

concentrations below 75 ppb, the ISA also combined the individual data from multiple studies of 

healthy adults exposed for 6.6 hours to 60 ppb O3 (Kim et al., 2011; Schelegle et al., 2009; 

Adams, 2006, 2002, 1998). Based on these data, the ISA reports that 10% of exposed subjects 

experienced FEV1 decrements of 10% or more (i.e., abnormal and large enough to be potentially 

adverse for people with pulmonary disease, based on past CASAC advice (section 3.1.3, 

above))41 (U.S. EPA, 2013, section 6.2.1.1). Consistent with these findings, recently developed 

 

36 Based on study population means. 

37 As noted above, for the 70 ppb exposure concentration Schelegle et al. (2009) reported that the actual mean 

exposure concentration was 72 ppb. 

38 As noted above, for the 60 ppb exposure concentration Schelegle et al. (2009) reported that the actual mean 

exposure concentration was 63 ppb. 

39 In an analysis of the Adams (2006) data for square-wave chamber exposures, even after removal of potential 

outliers, Brown et al. (2008) reported the average effect on FEV1 at 60 ppb to be statistically significant (p < 0.002) 

using several common statistical tests (U.S. EPA, 2013, section 6.2.1.1) (section 3.1.2.1, above). 

40 Adams (2006) reported increased respiratory symptoms during a 6.6 hour exposure protocol with an average O3

exposure concentration of 60 ppb. The increase in symptoms was reported to be statistically different from initial 

respiratory symptoms, though not statistically different from filtered air controls. 41 As noted above (section 3.1.3), CASAC has previously stated that “[a] 10% decrement in FEV1 can lead to 

respiratory symptoms, especially in individuals with pre-existing pulmonary or cardiac disease. For example, people 

with chronic obstructive pulmonary disease have decreased ventilatory reserve (i.e., decreased baseline FEV1) such 

that a ≥10% decrement could lead to moderate to severe respiratory symptoms” (Samet, 2011). 
3-59 

empirical models predict that the onset of O3-induced FEV1 decrements in healthy adults occurs 

following exposures to 60 ppb O3 for 4 to 5 hours while at moderate, intermittent exertion 

(Schelegle et al., 2012), and that 9% of healthy adults exposed to 60 ppb O3 for 6.6 hours would 

experience FEV1 decrements greater than or equal to 10% (McDonnell et al., 2012) (U.S. EPA, 

2013, section 6.2.1.1; section 3.1.2.1, above). When the evidence for O3-induced lung function 

decrements was taken together, the ISA concluded that (1) “mean FEV1 is clearly decreased by 

6.6-h exposures to 60 ppb O3 and higher concentrations in subjects performing moderate 

exercise” (U.S. EPA, 2013, p. 6-9) and (2) although group mean decrements following exposures 

to 60 ppb O3 are biologically small, “a considerable fraction of exposed individuals experience 

clinically meaningful decrements in lung function” (U.S. EPA, 2013, p. 6-20). 

In considering the specific question above, controlled human exposure studies have 

reported decreased lung function, increased airway inflammation, and increased respiratory 

symptoms in healthy adults following exposures to O3 concentrations below 75 ppb. Such 

impairments in respiratory function have the potential to be adverse, based on ATS guidelines 

for adversity and based on previous advice from CASAC (section 3.1.3, above). In addition, if 

they become serious enough, these respiratory effects could lead to the types of clearly adverse 

effects commonly reported in O3 epidemiologic studies (e.g., respiratory emergency department 

visits, hospital admissions). Therefore, following exposures to O3 concentrations lower than 75 

ppb, controlled human exposure studies have reported respiratory effects that could be adverse in 

some individuals, particularly if experienced by members of at-risk populations (e.g., asthmatics, 

children).42

In further considering effects following exposures to O3 concentrations below 75 ppb, we 

also note that the ISA highlights some epidemiologic panel studies for the extent to which 

monitored ambient O3 concentrations reflect exposure concentrations in their study populations 

(U.S. EPA, 2013, section 6.2.1.2). Specifically, Table 3-2 below includes O3 panel studies that 

have evaluated associations with lung function decrements for O3 concentrations at or below 75 

ppb, and that measured O3 concentrations with monitors located in the areas where study 

subjects were active (e.g., on site at summer camps or in locations where exercise took place) 

(U.S. EPA, 2013, section 6.2.1.2 and Table 6-6). Epidemiologic panel studies have evaluated a 

wider range of populations and lifestages than controlled human exposure studies of O3

concentrations below 75 ppb (e.g., including children). 

 

42 These effects were reported in healthy individuals. Consistent with past CASAC advice (Samet, 2011), and 

evidence in the ISA (U.S. EPA, 2013, p. 6-77), it is a reasonable inference that the effects would be greater in 

magnitude and potential severity for at-risk groups. See National Environmental Development Ass’n Clean Air 

Project v. EPA, 686 F. 3d 803, 811 (D.C. Cir. (2012) (making this point). 
3-60 

Table 3-2. Panel studies of lung function decrements with analyses restricted to O3

concentrations below 75 ppb. 

Although these studies report health effect associations for different averaging times, and it is not 

clear the extent to which specific O3 exposure conditions (i.e., concentrations, durations of 

exposure, degrees of activity) were responsible for eliciting reported decrements, they are 

consistent with the findings of the controlled human exposure studies discussed above. 

Specifically, the epidemiologic panel studies in Table 3-2 indicate O3-associated lung function 

decrements when on-site monitored concentrations (ranging from minutes to hours) were below 

75 ppb, with the evidence becoming less consistent at lower O3 concentrations. 

3.1.4.2 Concentrations in Epidemiologic Studies – Short-term Metrics 

We next consider distributions of ambient O3 concentrations in locations where 

epidemiologic studies have evaluated O3-associated hospital admissions, emergency department 

visits, and/or mortality. When considering epidemiologic studies within the context of the current 

standard, we emphasize those studies conducted in the U.S. and Canada. Such studies reflect air 

quality and exposure patterns that are likely more typical of the U.S. population than the air 

quality and exposure patterns reflected in studies conducted outside the U.S. and Canada (section 

1.3.1.2, above).43 We also emphasize studies reporting associations with effects judged in the 

ISA to be robust to confounding by other factors, including co-occurring air pollutants. In 

addition to these factors, we consider the statistical precision of study results, the extent to which 

 

43 Nonetheless, we recognize the importance of all studies, including international studies, in the ISA’s assessment 

of the weight of the evidence that informs causality determinations. 

Study Population O3 Concentrations Statistically Significant Association 

with Lung Function Decrements 

Spektor et al. 

(1988a) 

Children at 

summer camp 

Restricted to 1-hour concentrations 

below 60 ppb Yes 

Chan and 

Wu (2005) Mail carriers Maximum 8-hour average was 65 ppb Yes 

Korrick et al. 

(1998) Adult hikers 2- to 12-hour average from 40 to 74 ppb during hikes Yes 

Brauer et al. 

(1996) Farm workers 

Restricted to 1-hour maximum 

below 40 ppb Yes 

Restricted to 1-hour maximum 

below 30 ppb No 

Brunekreef 

et al. (1994) 

Exercising 

adults 

Restricted to 10-minute to 2.4-hour 

averages below 61 ppb No 

Restricted to 10-minute to 2.4-hour 

averages below 51 ppb No 

Restricted to 10-minute to 2.4-hour 

averages below 41 ppb No 
3-61 

studies report associations in at-risk populations, and the extent to which the biological 

plausibility of associations at various ambient O3 concentrations is supported by controlled 

human exposure and/or animal toxicological studies. These considerations help inform the range 

of ambient O3 concentrations over which we have the most confidence in O3-associated health 

effects, and the range of concentrations over which our confidence in such associations is 

appreciably lower. We place particular emphasis on characterizing those portions of distributions 

of ambient O3 concentrations likely to meet the current standard. 

In our consideration of these issues, we first address the following question: 

 To what extent have U.S. and Canadian epidemiologic studies reported 

associations with mortality or morbidity in locations that would have met the 

current O3 standard during the study period? 

Addressing this question can provide important insights into the extent to which O3-health effect 

associations are present for distributions of ambient O3 concentrations that would be allowed by 

the current standard. To the extent O3 health effect associations are reported in study areas that 

would have met the current standard, we have greater confidence that the current standard could 

allow the clearly adverse O3-associated effects indicated by those studies (e.g., mortality, 

hospital admissions, emergency department visits).44 

Epidemiologic studies evaluate statistical associations between variation in the incidence 

of health outcomes and variation in ambient O3 concentrations. In many of the O3 epidemiologic 

studies assessed in the ISA, ambient concentrations are averaged across multiple monitors within 

study areas, and in some cases over multiple days. These averages are used as surrogates for the 

spatial and temporal patterns of O3 exposures in study populations. In this PA, we refer to these 

averaged concentrations as “area-wide” O3 concentrations. 

The area-wide concentrations reported in many epidemiologic studies do not identify the 

actual O3 exposures that may be eliciting the observed health outcomes. Thus, in considering 

epidemiologic studies of mortality and morbidity, we are not drawing conclusions regarding 

single short-duration O3 concentrations in ambient air that, alone, are eliciting the reported health 

outcomes. Rather, our focus in this section is to consider what these studies convey regarding the 

extent to which health effects may be occurring (i.e., as indicated by associations) under air 

quality conditions meeting the current standard. 

 

44 See ATA III, 283 F.3d at 370 (EPA justified in revising NAAQS when health effect associations are observed at 

levels allowed by the NAAQS). 
3-62 

In order to facilitate consideration of the question above, we have identified U.S. and 

Canadian studies of respiratory hospital admissions, respiratory emergency department visits,45

and mortality (total, respiratory, cardiovascular) from the ISA (studies identified from U.S. EPA, 

2013, Tables 6-28, 6-42, and 6-53, and section 6.2.8) (Appendix 3B). For each monitor in the 

areas evaluated by these studies, we have identified the 3-year averages of the annual 4th highest 

daily maximum 8-hour O3 concentrations (Appendix 3B).46 To provide perspective on whether 

study cities would have met or violated the current O3 NAAQS during the study period, these O3

concentrations were compared to the level of the current standard. Based on this approach, a 

study city was judged to have met the current standard during the study period if all of the 3-year 

averages of annual 4th highest 8-hour O3 concentrations in that area were at or below 75 ppb. 

Based on these analyses, the large majority of epidemiologic study areas evaluated would 

have violated the current standard during study periods (Appendix 3B). Table 3-3 below 

highlights the subset of U.S. and Canadian studies that evaluated O3 health effect associations in 

locations that would have met the current standard during study periods. This includes a U.S. 

single-city study that would have met the current standard over the entire study period (Mar and 

Koenig, 2009) and four Canadian multicity studies for which the majority of study cities would 

have met the current standard over the entire study periods (Cakmak et al., 2006; Dales et al., 

2006; Katsouyanni et al., 2009; Stieb et al., 2009).47 

 

45 Given the inconsistency in results across cardiovascular morbidity studies (U.S. EPA, 2013, section 6.3.2.9), our 

consideration of the morbidity evidence in this section focuses on studies of respiratory hospital admissions and 

emergency department visits. 

46 These concentrations are referred to as “design values.” A design value is a statistic that is calculated at individual 

monitors and based on 3 consecutive years of data collected from that site. In the case of O3, the design value for a 

monitor is based on the 3-year average of the annual 4th highest daily maximum 8-hour O3 concentration in parts 

per billion (ppb). For U.S. study areas, we used EPA’s Air Quality System (AQS) 

(http://www.epa.gov/ttn/airs/airsaqs/) to identify design values. For Canadian study areas, we used publically 

available air quality data from the Environment Canada National Air Pollution Surveillance Network 

(http://www.etc-cte.ec.gc.ca/napsdata/main.aspx). We followed the data handling protocols for calculating design 

values as detailed in 40 CFR Part 50, Appendix P. 

47 In addition, a study by Vedal et al. (2003) was included in the 2006 CD (U.S. EPA, 2006). This study reported 

positive and statistically significant associations with mortality in Vancouver during a time period when the study 

area would have met the current standard (U.S. EPA, 2007). This study was not highlighted in the ISA in the current 

review (U.S. EPA, 2013). 
3-63 

Table 3-3. U.S. and Canadian epidemiologic studies reporting O3 health effect 

associations in locations that would have met the current standard during 

study periods. 

Authors Study Results Cities 

Number of cities meeting 

the current standard over 

entire study period 

Cakmak et al. 

(2006) 

Positive and statistically significant association 

with respiratory hospital admissions 

10 

Canadian 

cities 

7 

Dales et al. 

(2006) 

Positive and statistically significant association 

with respiratory hospital admissions 

11 

Canadian 

cities 

7 

Katsouyanni et 

al. (2009) 

Positive and statistically significant associations 

with respiratory hospital admissions 

12 

Canadian 

cities 

10 

Katsouyanni et 

al. (2009) 

Positive and statistically significant associations 

with all-cause and cardiovascular mortality48

12 

Canadian 

cities 

8 

Mar and 

Koenig (2009) 

Positive and statistically significant associations 

with asthma emergency department visits in 

children (< 18 years) and adults (> 18 years) 

Seattle 1 

Stieb et al. 

(2009) 

Positive and statistically significant association 

with asthma emergency department visits 

7 Canadian 

cities 5 

As illustrated in Table 3-3, Mar and Koenig reported health effect associations with 

asthma emergency department visits in a location that would have met the current standard over 

the entire study period. This analysis indicates that the current standard would allow the 

distribution of ambient O3 concentrations that provided the basis for reported associations with 

respiratory emergency department visits. 

In addition, four multicity studies reported associations with mortality or morbidity when 

the majority of study locations would have met the current standard over the entire study periods. 

Thus, the current standard would allow the majority of the distributions of ambient O3

concentrations that provided the basis for positive and statistically significant associations with 

mortality or morbidity. Our interpretation of these results is complicated by uncertainties in the 

extent to which multicity effect estimates (i.e., which are based on combining estimates from 

multiple study locations) can be attributed to ambient O3 in the subset of locations that would 

have met the current standard, versus O3 in the smaller number of locations that would have 

violated the standard. While there is uncertainty in ascribing the multicity effect estimates 

reported in these Canadian studies to ambient concentrations that would have met the current 

 

48 Katsouyanni et al. (2009) report a positive and statistically significant association with cardiovascular mortality 

for people aged 75 years or older. 
3-64 

standard, the information in Table 3-3 suggests that reported multicity effect estimates are 

largely influenced by locations meeting the current standard (i.e., given that most study areas 

would have met this standard). Together, these U.S. and Canadian epidemiologic studies suggest 

a relatively high degree of confidence in the presence of associations with mortality and 

morbidity for ambient O3 concentrations meeting the current standard. 

We next consider the extent to which additional epidemiologic studies of mortality or 

morbidity, specifically those conducted in locations that violated the current standard, can also 

inform our consideration of adequacy of the current standard. In doing so, we note that health 

effect associations reported in epidemiologic studies are influenced by the full distributions of 

ambient O3 concentrations, including concentrations below the level of the current standard. We 

focus on studies that have explicitly characterized such O3 health effect associations, including 

confidence in those associations, for various portions of distributions of ambient O3

concentrations. In doing so, we consider the following question: 

 To what extent do analyses from epidemiologic studies indicate confidence in health 

effect associations over distributions of ambient O3 concentrations, including at 

concentrations lower than previously identified or below the current standard? 

 We first focus on those studies that have reported confidence intervals around 

concentration-response functions over distributions of ambient O3 concentrations. Confidence 

intervals around concentration-response functions can provide insights into the range of ambient 

concentrations over which the study indicates the most confidence in the reported health effect 

associations (i.e., where confidence intervals are narrowest), and into the range of ambient 

concentrations below which the study indicates that uncertainty in the nature of such associations 

becomes notably greater (i.e., where confidence intervals become markedly wider). The 

concentrations below which confidence intervals become markedly wider in such analyses are 

intrinsically related to data density, and do not necessarily indicate the absence of an association. 

 The ISA identifies several epidemiologic studies that have reported confidence intervals 

around concentration-response functions in U.S. cities. The ISA concludes that studies generally 

indicate a linear concentration-response relationship “across the range of 8-h max and 24 h avg 

O3 concentrations most commonly observed in the U.S. during the O3 season” and that “there is 

less certainty in the shape of the C-R curve at the lower end of the distribution of O3

concentrations” (U.S. EPA, 2013, pp. 2-32 to 2-34). In characterizing the O3 concentrations 

below which such certainty decreases, the ISA discusses area-wide O3 concentrations as low as 

20 ppb and as high as 40 ppb (U.S. EPA, 2013, section 2.5.4.4). 

Consistent with these conclusions, the range of ambient concentrations over which the 

evidence indicates the most certainty in concentration-response relationships can vary across 
3-65 

studies. Such variation is likely due at least in part to differences in the O3 metrics evaluated and 

differences in the distributions of ambient concentrations and health events. Thus, although 

consideration of confidence intervals around concentration-response functions can provide 

valuable insights into the ranges of ambient concentrations over which studies indicate the most 

confidence in reported health effect associations, there are limitations in the extent to which 

these analyses can be generalized across O3 metrics, study locations, study populations, and 

health endpoints. 

The ISA emphasizes two U.S. single-city studies that have reported confidence intervals 

around concentration-response functions (Silverman and Ito, 2010; Strickland et al., 2010). 

These studies, and their associated O3 air quality, are discussed below. 

Silverman and Ito (2010) evaluated associations between 2-day rolling average O3

concentrations49 and asthma hospital admissions in New York City from 1999 to 2006 (a time 

period when the study area would have violated the current standard, Appendix 3B). As part of 

their analysis, the authors evaluated the shape of the concentration-response relationship for O3

using a co-pollutant model that included PM2.5 (reprinted in Figure 3-4, below). Based on their 

analyses, Silverman and Ito (2010) concluded a linear relationship between O3 and hospital 

admissions is a reasonable approximation of the concentration-response function throughout 

much of the range of ambient O3 concentrations. Based on visual inspection of Figure 3-4 below 

(Figure 3 from published study), we note that confidence in the reported concentration-response 

relationship is highest for area-wide average O3 concentrations around 40 ppb (i.e., near the 

reported median of 41 ppb), and decreases notably for concentrations at and below about 20 ppb. 

 

 

49 2-day rolling averages of daily maximum 8-hour O3 concentrations were calculated throughout the study period, 

averaged across study monitors. 
3-66 

Figure 3-4. Concentration-response function for asthma hospital admissions over the 

distribution of area-wide averaged O3 concentrations (adapted from Silverman 

and Ito, 2010).50 

 In considering the concentration-response function presented by Silverman and Ito (2010) 

within the context of the adequacy of the current standard, we recognize that true design values 

cannot be identified for the subsets of air quality data contributing to various portions of the 

concentration-response function.51 Therefore, to use this analysis to inform our consideration of 

the adequacy of the current standard we evaluate the extent to which the concentration-response 

function indicates a relatively high degree of confidence in the reported health effect association 

on days when all monitored 8-hour O3 concentrations were below 75 ppb (Table 3-4, below). 

This approach can provide insight into the extent to which the reported O3 health effect 

association is present when all monitored O3 concentrations are below the level of the current 

standard. 

 Based on the information in Table 3-4 below, when 2-day averaged O3 concentrations 

ranged from 36 to 45 ppb (i.e., around the median, where confidence intervals are narrowest), 

there were 3 days (out of 432) with at least one monitor recording a daily maximum 8-hour O3

concentration above the level of the current standard (approximately 0.7% of days). When 2-day 

 

50This figure was also reprinted in the ISA (U.S. EPA, 2013; Figure 6-16). 

51As discussed above, O3 design values are calculated using all data available from a monitor. Relative Risk

O3 Concentration (ppb)
3-67 

averaged O3 concentrations ranged from 26 to 45 ppb (i.e., extending to concentrations below the 

median, but still above the concentrations where confidence intervals widen notably), there were 

4 days (out of 816) with at least one monitor recording a daily maximum 8-hour O3

concentration above the level of the current standard (approximately 0.5% of days). Thus, on 

over 99% of the days when area-wide “averaged” O3 concentrations were between 26 and 45 

ppb, the highest daily maximum 8-hour O3 concentrations were below 75 ppb. For comparison, 

the annual 4th highest daily maximum 8-hour O3 concentration generally corresponds to the 98th

or 99th percentile of the seasonal distribution, depending on the length of the O3 season. 

Table 3-4. Distributions of daily 8-hour maximum ozone concentrations from highest 

monitors over range of 2-day moving averages from composite monitors (for 

study area evaluated by Silverman and Ito, 2010) 

 In a separate study, Strickland et al. (2010) evaluated associations between 3-day rolling 

average O3 concentrations52 and asthma hospital admissions in Atlanta during the warm season 

from 1994 to 2004 (a time period when the study area would have violated the current standard, 

Appendix 3B). As part of this analysis, Strickland et al. (2010) evaluated the concentrationresponse relationship for O3 and pediatric asthma emergency department visits. The authors 

reported the shape of the concentration-response function to be approximately linear with no 

evidence of a threshold when 3-day averaged daily maximum 8-hour O3 concentrations were 

approximately 30 to 80 ppb (Figure 3-5 below and U.S. EPA, 2013, Figure 6-18). Figure 3-5 

below illustrates that the confidence intervals around the concentration-response function are 

 

52 Three-day rolling averages of population-weighted daily maximum 8-hour O3 concentrations were calculated 

throughout the study period (Strickland et al., 2010). 
3-68 

narrowest around the study mean (i.e., 55 ppb), and that these confidence intervals do not widen 

notably for “averaged” O3 concentrations as low as about 30 ppb. 

Figure 3-5. Concentration-response function for pediatric asthma emergency department 

visits over the distribution of averaged, population-weighted 8-hour O3 

concentrations (reprinted from Strickland et al., 2010).53

Similar to the study by Silverman and Ito (2010), we consider the extent to which the 

reported concentration-response function indicates a relatively high degree of confidence in 

health effect associations on days when all monitored 8-hour O3 concentrations are below 75 ppb 

(Table 3-5, below).54 In considering the information presented in Table 3-5, we first note that 

when 3-day averaged O3 concentrations were in the range of the mean (i.e., 51 to 60 ppb), there 

were 77 days (out of 516; 14.9%) with at least one monitor recording a daily maximum 8-hour 

O3 concentration above the level of the current standard. In contrast, during the 519 days when 

averaged O3 concentrations were in the lower portion of the distribution where study authors 

indicate relatively high confidence in the reported concentration-response relationship (i.e., 

between 31 and 45 ppb), there were 4 days with at least one monitor in the study area measuring 

a daily maximum 8-hour O3 concentration greater than 75 ppb (approximately 0.8% of days). 

 

53 This figure was also reprinted in the ISA (U.S. EPA, 2013; Figure 6-18). 

54 The study by Strickland et al. (2010) used five monitors. For our evaluation of highest daily maximum 8-hour 

concentrations (i.e., from the individual monitor recording the highest such concentration), we obtained information 

from the four of these study area monitors that report data to AQS (Appendix 3B). 
3-69 

Thus, on over 99% of the days when “averaged” O3 concentrations were between 31 and 45 ppb, 

all monitors measured daily maximum 8-hour O3 concentrations below 75 ppb. 

Table 3-5. Distribution of daily 8-hour maximum ozone concentrations from highest 

monitors over range of 3-day moving averages of population-weighted 

concentrations (for study area evaluated by Strickland et al., 2010) 

In summary, analyses of air quality data from the study locations evaluated by Silverman 

and Ito (2010) and Strickland et al. (2010) indicate a relatively high degree of confidence in 

reported statistical associations with respiratory health outcomes on days when virtually all 

monitored 8-hour O3 concentrations were 75 ppb or below. Though these analyses do not 

identify true design values, the presence of O3-associated respiratory effects on such days 

provides insight into the types of health effects that could occur in locations with maximum 

ambient O3 concentrations below the level of the current standard. 

We next consider the following question: 

 To what extent are there important uncertainties in analyses of confidence in 

concentration-response functions? 

There are several important uncertainties that are specifically related to our analyses of 

distributions of O3 air quality in the study locations evaluated by Silverman and Ito (2010) and 

Strickland et al. (2010). Although these studies report health effect associations with two-day 

(Silverman and Ito) and three-day (Strickland) averages of daily O3 concentrations, it is possible 

that the respiratory morbidity effects reported in these studies were also at least partly 

attributable to the days immediately preceding these two- and three-day periods. In support of 

this possibility, Strickland et al. reported positive and statistically significant associations with 

emergency department visits for multiple lag periods, including lag periods exceeding three days. 
3-70 

Our analysis of highest monitored concentrations focuses on two- and three- day periods, as used 

in the published study to generate concentration-response functions. This could have important 

implications for our interpretation of the reported concentration-response functions if a 2-day 

period with no monitors measuring 8-hour concentrations at or above 75 ppb is immediately 

preceded by one or more days with monitors that do exceed 75 ppb. Although we do not know 

the extent to which O3 concentrations on a larger number of days could have contributed to 

reported health effect associations, we note this as a potentially important uncertainty in our 

consideration of concentration-response functions within the context of the current standard. 

In addition, an important uncertainty that applies to epidemiologic studies in general is 

the extent to which reported health effects are caused by exposures to O3 itself, as opposed to 

other factors such as co-occurring pollutants or other pollutant mixtures. Although both of the 

studies evaluated above reported health effect associations in co-pollutant models, this 

uncertainty becomes an increasingly important consideration as health effect associations are 

evaluated at lower ambient O3 concentrations (i.e., presumably corresponding to lower exposure 

concentrations). 

One approach to considering the potential importance of this uncertainty in 

epidemiologic studies is to evaluate the extent to which there is coherence with the results of 

experimental studies (i.e., in which the study design dictates that exposures to O3 itself are 

responsible for reported effects). Therefore, in further considering uncertainties associated with 

the above air quality analyses for the study areas evaluated by Silverman and Ito (2010) and 

Strickland et al. (2010), we evaluate the following question: 

 To what extent is there coherence between evidence from controlled human 

exposure studies and epidemiologic studies supporting the occurrence of O3-

attributable respiratory effects when daily maximum 8-hour ambient O3 

concentrations are at or below 75 ppb? 

As summarized above and as discussed in the ISA (U.S. EPA, 2013, section 6.2), 

controlled human exposure studies demonstrate the occurrence of respiratory effects in an 

appreciable percentage of healthy adults following single short-term exposures to O3

concentrations as low as 60 ppb. As O3 exposure concentrations exceed 60 ppb: 1) effects in 

healthy adults become larger and more serious; 2) a broader range of effects are observed in a 

greater percentage of exposed individuals; and 3) effects are reported more consistently across 

studies. In addition, exposure concentrations below 60 ppb could potentially result in respiratory 

effects, particularly in at-risk populations such as children and asthmatics. Thus, as the potential 

increases for portions of epidemiologic study populations to be exposed to O3 concentrations 

approaching or exceeding 60 ppb, our confidence increases that reported respiratory health 

effects could be caused by exposures to the ambient O3 concentrations present in study locations. 
3-71 

 As discussed above, for the study by Silverman and Ito (2010), 26 to 45 ppb represents 

the lower end of the range of “averaged” concentrations over which the study indicates a 

relatively high degree of confidence in the statistical association with respiratory hospital 

admissions (and for which virtually all monitored concentrations were 75 ppb or below). As 

averaged concentrations increase from 26 to 45 ppb, the number of days with maximum 

monitored concentrations approaching or exceeding 60 ppb increases (Table 3-4, above).55 For 

example, of the 178 days with area-wide average concentrations from 26 to 30 ppb, only about 

5% had monitors recording ambient concentrations of 50 ppb or greater and about 2% had 

monitors recording concentrations of 60 ppb or greater. In contrast, of the 196 days with areawide average concentrations from 41 to 45 ppb, about half had one or more monitors recording 

ambient concentrations above 50 ppb and about 25% had monitors recording concentrations at or 

above 60 ppb. On a small number of these days, at least one monitored concentration exceeded 

70 or 80 ppb. Thus as averaged concentrations approach 45 ppb there is an increasing likelihood 

that at least some portion of the Silverman and Ito study population could have been exposed to 

O3 concentrations near or above those shown to cause respiratory effects in healthy adults. If 

these effects become serious enough (e.g., in people with asthma) they could lead to the 

respiratory-related hospital admissions reported in the study. This analysis is consistent with the 

occurrence of O3-attributable respiratory hospital admissions, even when virtually all monitored 

concentrations were below the level of the current standard. Similar results were obtained for the 

study by Strickland et al. (2010) (Table 3-5, above). 

In further evaluating O3 concentration-response relationships within the context of the 

adequacy of the current standard, we note that some epidemiologic studies report health effect 

associations for air quality subsets restricted to ambient pollutant concentrations below one or 

more predetermined cut points. Such “cut point” analyses can provide information on the 

magnitude and statistical precision of effect estimates for defined distributions of ambient 

concentrations, which may in some cases include distributions that would meet the current 

standard. Therefore, we next consider the following question: 

 To what extent do cut-point analyses from epidemiologic studies report health effect 

associations at ambient O3 concentrations lower than previously identified or that 

would likely meet the current standard?

By considering the magnitude and statistical significance of effect estimates for restricted 

air quality distributions, cut-point analyses can provide insight into the extent to which health 

 

55 Though, as noted above, the epidemiologic studies by Silverman and Ito (2010) and Strickland et al. (2010) do 

not provide information on the extent to which reported health effects result from exposures to any specific O3

concentrations. 
3-72 

effect associations are driven by ambient concentrations above the cut point, versus 

concentrations below the cut point. For studies that evaluate multiple cut points, these analyses 

can provide insights into the magnitude and statistical precision of health effect associations for 

different portions of the distribution of ambient concentrations, including insights into the 

ambient concentrations below which uncertainty in reported associations becomes notably 

greater. As with analyses of concentration-response functions, discussed above, the cut points 

below which confidence intervals become notably wider depend in large part on data density.56 

In the U.S. multicity study by Bell et al. (2006), study authors used the NMMAPS data 

set to evaluate associations between 2-day rolling average O3 concentrations57 and total (nonaccidental) mortality in 98 U.S. cities from 1987 to 2000. Based on the full distributions of 

ambient O3 concentrations in study cities, the large majority of the NMMAPS cities would have 

violated the current standard during the study period (Appendix 3B). However, Bell et al. (2006) 

also reported health effect associations in a series of cut-point analyses, with effect estimates 

based only on the subsets of days contributing to “averaged” O3 concentrations below cut points 

ranging from 5 to 60 ppb (see Figure 2 in Bell et al., 2006). The lowest cut-point for which the 

association between O3 and mortality was reported to be statistically significant was 30 ppb 

(based on visual inspection of Figure 2 in the published study). As with the studies by Silverman 

and Ito (2010) and Strickland et al. (2010), discussed above, we consider what these cut point 

analyses indicate with regard to the potential for health effect associations to extend to ambient 

O3 concentrations likely to be allowed by the current O3 NAAQS. 

We attempted to recreate the subsets of air quality data used in the cut point analyses 

presented by Bell et al. (2006). In doing so, we applied the criteria described in the published 

study to generate air quality subsets corresponding to those defined by the cut points evaluated 

by study authors.58 From the days with averaged O3 concentrations below each cut point, we 

identified 3-year averages of annual 4th highest daily maximum 8-hour O3 concentrations in each 

study area. We then compared these 4th highest O3 concentrations to the level of the current 

standard in order to provide insight into the extent to which the air quality distributions included 

in various cut point analyses would likely have met the current standard. 

 

56 As such, these analyses provide insight into the ambient concentrations below which the available air quality 

information becomes too sparse to support conclusions about the nature of concentration-response relationships, 

with a high degree of confidence. 

57 Two-day rolling averages of 24-hour average O3 concentrations were calculated throughout the study period. This 

calculation was done across study monitors in study cities with multiple monitors. 

58 We were unable to obtain the air quality data used to generate the cut-point analyses in the study published by 

Bell et al. (2006). Therefore, we generated 2-day averages of 24-hour O3 concentrations in study locations using the 

air quality data available in AQS, combined with the published description of study area definitions. In doing so, we 

did not recreate the trimmed means used by Bell. As discussed below, this represents an important uncertainty in our 

analysis. 
3-73 

We particularly focus on the lowest cut-point for which the association between O3 and 

mortality was reported in this study to be statistically significant (i.e., 30 ppb, as noted above). 

Based on the O3 air quality concentrations that met the criteria for inclusion in the 30 ppb cut 

point analysis, 95% of study areas had 3-year averages of annual 4th highest daily maximum 8-

hour O3 concentration at or below 75 ppb over the entire study period. For the 35 ppb cut point, 

which also resulted in a statistically significant association with mortality, 68% of study areas 

had 3-year averages of annual 4th highest daily maximum 8-hour O3 concentration at or below 75 

ppb. This suggests that the large majority of air quality distributions that provided the basis for 

positive and statistically significant associations with mortality (i.e., for the 30 and 35 ppb cut 

points) would likely have met the current O3 standard. For higher cut points, all of which also 

resulted in statistically significant associations with mortality, the majority of study cities had 3-

year averages of annual 4th highest daily maximum 8-hour concentrations greater than 75 ppb. 
3-74 

Table 3-6. Number of study cities with 4th highest daily maximum 8-hour concentrations 

greater than 75 ppb, for various cut-point analyses presented in Bell et al. 

(2006) 

Cut-point for 2-day moving average across monitors and cities (24-hour average) 

25 30 35 40 45 50 55 60 All 

Number (%) of 

Cities with 4th 

highest >75 (any 

3-yr period; 

1987-2000) 

0 (0%) 5 (5%) 31 

(32%) 

70 

(71%) 

86 

(88%) 

88 

(90%) 

92 

(94%) 

92 

(94%) 

92 

(94%) 

In addition to the uncertainties noted above for our analysis of the single-city studies by 

Silverman and Ito (2010) and Strickland et al. (2010) (e.g., attributing effects specifically to air 

quality included in various subsets), an important uncertainty related to this analysis is that we 

were unable to obtain the air quality data used to generate the cut-point analyses in the study 

published by Bell et al. (2006). Therefore, as noted above, we generated 2-day averages of 24-

hour O3 concentrations in study locations using the air quality data available in AQS, combined 

with the published description of study area definitions. In doing so, we did not recreate the 

trimmed means used by Bell. An important uncertainty in this approach is the extent to which we 

were able to appropriately recreate the cut-point analyses in the published study. 

The ISA also notes important uncertainties inherent in multicity studies that evaluate the 

potential for thresholds to exist, as was done in the study by Bell et al. (2006). Specifically, the 

ISA highlights the regional heterogeneity in O3 health effect associations as a factor that could 

obscure the presence of thresholds, should they exist, in multicity studies (U.S. EPA, 2013, 

sections 2.5.4.4 and 2.5.4.5). The ISA notes that community characteristics (e.g., activity 

patterns, housing type, age distribution, prevalence of air conditioning) could be important 

contributors to reported regional heterogeneity (U.S. EPA, 2013, section 2.5.4.5). Given this 

heterogeneity, the ISA concludes that “a national or combined analysis may not be appropriate to 

identify whether a threshold exists in the O3-mortality C-R relationship” (U.S. EPA, 2013, p. 2-

33). This represents an important source of uncertainty when characterizing our confidence in 

reported concentration-response relationships over distributions of ambient O3 concentrations, 

based on multicity studies. This uncertainty becomes increasingly important when interpreting 

concentration-response relationships at lower ambient O3 concentrations, particularly those 

concentrations corresponding to portions of distributions where data density decreases notably. 

3.1.4.3 Concentrations in Epidemiologic Studies – “Long-term” Metrics 

We next consider the extent to which epidemiologic studies employing longer-term 

ambient O3 concentration metrics inform our understanding of the air quality conditions 
3-75 

associated with O3-attributable health effects, and specifically inform consideration of the extent 

to which such effects could occur under air quality conditions meeting the current standard. 

Unlike for the studies of short-term O3 discussed above, the available U.S. and Canadian 

epidemiologic studies evaluating long-term ambient O3 concentration metrics have not been 

conducted in locations likely to have met the current 8-hour O3 standard during the study period 

(Appendix 3B). Therefore, although these studies contribute to our understanding of health 

effects associated with long-term or repeated exposures to ambient O3 (as summarized in section 

3.1.2 above), consideration of study area design values does not inform our consideration of the 

extent to which those health effects may be occurring in locations that met the current standard. 

In further considering epidemiologic studies of long-term O3 concentrations, we also 

evaluate the extent to which concentration-response functions, including associated confidence 

intervals, have been characterized for distributions of ambient O3, and what those functions can 

tell us about health effect associations for O3 concentrations likely to be allowed by the current 

standard. Specifically, we consider the following question: 

 To what extent do confidence intervals around concentration-response functions 

indicate O3-associated health outcomes at ambient concentrations meeting the 

current O3 standard? 

The ISA identifies a single epidemiologic study reporting confidence intervals around a 

concentration-response function for “long-term” O3 concentrations and respiratory mortality 

(Jerrett et al., 2009; U.S. EPA, 2013, sections 7.2.7, 7.2.8 and 7.7). Jerrett et al. (2009) reported 

that when seasonal averages of 1-hour daily maximum O3 concentrations59 ranged from 33 to 

104 ppb, there was no statistical deviation from a linear concentration-response relationship 

between O3 and respiratory mortality across 96 U.S. cities (U.S. EPA, 2013, section 7.7). 

However, the authors reported “limited evidence” for an effect threshold at an O3 concentration 

of 56 ppb (p=0.06).60 Visual inspection of this concentration-response function (Figure 3-6) 

confirms the possibility of an inflection point just below 60 ppb, which is close to the median 

concentration across cities (i.e., 57 ppb). 

 

59 Jerrett et al. (2009) evaluated the April to September averages of 1-hour daily maximum O3 concentrations across 

96 U.S. metropolitan areas from 1977- 2000. In urban areas with multiple monitors, April to September 1-hour daily 

maximum concentrations from each individual monitor were averaged. This step was repeated for each year in the 

study period. Finally, each yearly averaged O3 concentrations was then averaged again to yield the single averaged 

1-hour daily maximum O3 concentration depicted on the x axis of Figure 3-6 below. 60 This issue is discussed further in section 3.2.3.2, below. 
3-76 

Figure 3-6. Exposure-Response relationship between risk of death from respiratory causes 

and ambient O3 concentration study metric (Jerrett et al., 2009). 

We consider the extent to which this concentration-response function indicates 

confidence in the reported health effect association at various ambient O3 concentrations. In 

identifying the concentrations over which we have the greatest confidence, we note the 

following: (1) most of the study cities had O3 concentrations above 53.1 ppb (i.e., the upper 

bound of the first quartile), accounting for approximately 72% of the respiratory deaths in the 

cohort (Table 2 in Jerrett et al. 2009); (2) confidence intervals widen notably for O3

concentrations in the first quartile (based on visual inspection of Figure 3-6); and (3) study 

authors noted limited evidence for a threshold at 56 ppb.61 In considering this information, we 

conclude that the analysis reported by Jerrett indicates a relatively high degree of confidence in 

the linear concentration-response function for “long-term” O3 concentrations at least as low as 56 

ppb, and notably decreased confidence in the linear function for concentrations at or below about 

53 ppb (i.e., the upper bound of the first quartile of O3 concentrations). 

Based on information in the published study (Figure 1 in Jerrett et al., 2009), we 

identified 72 of the 96 study cities as having ambient O3 concentrations in the highest three 

quartiles (Appendix 3B). As noted above, these 72 cities account for approximately 72% of the 

respiratory deaths in the cohort (Table 2 in Jerrett et al. 2009). Of these 72 cities, 71 had 3-year 

averages of annual 4th highest daily maximum 8-hour O3 concentrations above 75 ppb (Appendix 

3B). Thus, the current 8-hour NAAQS would have been violated during the study period in 

virtually all of the study cities that contribute to the range of long-term O3 concentrations over 

which we have the greatest confidence in the reported relationship with respiratory mortality. 

 

61 The ISA does not reach conclusions regarding the potential for a threshold in the association between “long-term” 

O3 concentrations and respiratory mortality. 
3-77 

Thus, while the study by Jerrett et al. (2009) contributes to our understanding of health effects 

associated with ambient O3 (as summarized in section 3.1.2 above), it is less informative 

regarding the extent to which those health effects may be occurring under air quality conditions 

allowed by the current standard. 

3.1.5 Public Health Implications 

In this section, we address the public health implications of O3 exposures with respect to 

the factors that put populations at increased risk from exposures (section 3.1.5.1), the size of atrisk populations (section 3.1.5.2), and the potential effects of averting behavior on reducing O3

exposures and associated health effects (section 3.1.5.3). Providing appropriate public health 

protection requires consideration of the factors that put populations at greater risk from O3

exposure. In order to estimate potential public health impacts, it is important to consider not only 

the adversity of the health effects, but also the populations at greater risk and potential behaviors 

that may reduce exposure. 

3.1.5.1 At-Risk Populations 

In this section we address the following question: 

 To what extent does the currently available scientific evidence expand our 

understanding of at-risk populations? 

The currently available evidence expands our understanding of populations that were 

identified to be at greater risk of O3-related health effects at the time of the last review (i.e., 

people who are active outdoors, people with lung disease, children and older adults and people 

with increased responsiveness to O3) and supports the identification of additional factors that 

may lead to increased risk (U.S. EPA, 2006, section 3.6.2; U.S. EPA, 2013, chapter 8). 

Populations and lifestages may be at greater risk for O3-related health effects due to factors that 

contribute to their susceptibility and/or vulnerability to ozone. The definitions of susceptibility 

and vulnerability have been found to vary across studies, but in most instances “susceptibility” 

refers to biological or intrinsic factors (e.g., lifestage, sex, preexisting disease/conditions) while 

“vulnerability” refers to non-biological or extrinsic factors (e.g., socioeconomic status [SES]) 

(U.S. EPA, 2013, p. 8-1). In some cases, the terms “at-risk” and “sensitive” have been used to 

encompass these concepts more generally. In the ISA and this PA, “at-risk” is the allencompassing term used to define groups with specific factors that increase their risk of 

O3-related health effects. Further discussion of at-risk populations can be found below. 
3-78 

There are multiple avenues by which groups may experience increased risk for O3-related 

health effects. A population or lifestage62 may exhibit greater effects than other populations or 

lifestages exposed to the same concentration or dose, or they may be at greater risk due to 

increased exposure to an air pollutant (e.g., time spent outdoors). A group with intrinsically 

increased risk would have some factor(s) that increases risk through a biological mechanism and, 

in general, would have a steeper concentration-risk relationship, compared to those not in the 

group. Factors that are often considered intrinsic include pre-existing asthma, genetic 

background, and lifestage. A group of people could also have extrinsically increased risk, which 

would be through an external, non-biological factor, such as socioeconomic status (SES) and 

diet. Some groups are at risk of increased internal dose at a given exposure concentration, for 

example, because of breathing patterns. This category would include people who work or 

exercise outdoors. Finally, there are those who might be placed at increased risk for experiencing 

greater exposures by being exposed to higher O3 concentrations. This would include, for 

example, groups of people with greater exposure to ambient O3 due to less availability or use of 

home air conditioners such that they are more likely to be in locations with open windows on 

high ozone days. Some groups may be at increased risk of O3-related health effects through a 

combination of factors. For example, children tend to spend more time outdoors when O3 levels 

are high, and at higher levels of activity than adults, which leads to increased exposure and dose, 

and they also have biological, or intrinsic, risk factors (e.g., their lungs are still developing) (U.S. 

EPA, 2013, Chapter 8). An at-risk population or lifestage is more likely to experience adverse 

health effects related to O3 exposures and/or, develop more severe effects from exposure than the 

general population. 

People with Specific Genetic Variants 

Overall, for variants in multiple genes there is adequate evidence for involvement in 

populations being more at-risk than others to the effects of O3 exposure on health (U.S. EPA, 

2013, section 8.1). Controlled human exposure and epidemiologic studies have reported evidence 

of O3-related increases in respiratory symptoms or decreases in lung function with variants 

including GSTM1, GSTP1, HMOX1, and NQO1. NQO1 deficient mice were found to be 

resistant to O3-induced AHR and inflammation, providing biological plausibility for results of 

studies in humans. Additionally, studies of rodents have identified a number of other genes that 

may affect O3-related health outcomes, including genes related to innate immune signaling and 

pro- and anti-inflammatory genes, which have not been investigated in human studies. 

People with Asthma 

 

62 Lifestages, which in this case includes childhood and older adulthood, are experienced by most people over the 

course of a lifetime, unlike other factors associated with at-risk populations. 
3-79 

Previous O3 AQCDs identified individuals with asthma as a population at increased risk 

of O3-related health effects. Multiple new epidemiologic studies included in the ISA have 

evaluated the potential for increased risk of O3-related health effects in people with asthma, 

including: lung function; symptoms; medication use; airway hyperresponsiveness (AHR); and 

airway inflammation (also measured as exhaled nitric oxide fraction, or FeNO). A study of 

lifeguards in Texas reported decreased lung function with short-term O3 exposure among both 

individuals with and without asthma, however, the decrease was greater among those with 

asthma (Thaller et al., 2008). A Mexican study of children ages 6-14 detected an association 

between short-term O3 exposure and wheeze, cough, and bronchodilator use among asthmatics 

but not non-asthmatics, although this may have been the result of a small non-asthmatic 

population (Escamilla-Nuñez et al., 2008). A study of modification by AHR (an obligate 

condition among asthmatics) reported greater short-term O3-associated decreases in lung 

function in elderly individuals with AHR, especially among those who were obese (Alexeeff et 

al., 2007). With respect to airway inflammation, in one study, a positive association was reported 

for airway inflammation among asthmatic children following short-term O3 exposure, but the 

observed association was similar in magnitude to that of non-asthmatics (Barraza-Villarreal et 

al., 2008). Similarly, another study of children in California reported an association between O3 

concentration and FeNO that persisted both among children with and without asthma as well as 

those with and without respiratory allergy (Berhane et al., 2011). Finally, Khatri et al. (2009) 

found no association between short-term O3 exposure and altered lung function for either 

asthmatic or non-asthmatic adults, but did note a decrease in lung function among individuals 

with allergies. 

New evidence for difference in effects among asthmatics has been observed in studies 

that examined the association between O3 exposure and altered lung function by asthma 

medication use. A study of children with asthma living in Detroit reported a greater association 

between short-term O3 and lung function for corticosteroid users compared with 

noncorticosteroid users (Lewis et al., 2005). Conversely, another study found decreased lung 

function among noncorticosteroid users compared to users, although in this study, a large 

proportion of non-users were considered to be persistent asthmatics (Hernández-Cadena et al., 

2009). Lung function was not related to short-term O3 exposure among corticosteroid users and 

non-users in a study taking place during the winter months in Canada (Liu et al., 2009). 

Additionally, a study of airway inflammation reported a counterintuitive inverse association with 

O3 of similar magnitude for all groups of corticosteroid users and non-users (Qian et al., 2009). 

Controlled human exposure studies that have examined the effects of O3 on adults with 

asthma and healthy controls are limited. Based on studies reviewed in the 1996 and 2006 O3

AQCDs, subjects with asthma appeared to be more sensitive to acute effects of O3 in terms of 
3-80 

FEV1 and inflammatory responses than healthy non-asthmatic subjects. For instance, Horstman 

et al. (1995) observed that mild-to-moderate asthmatics, on average, experienced double the 

O3-induced FEV1 decrement of healthy subjects (19% versus 10%, respectively, p = 0.04). 

Moreover, a statistically significant positive correlation between FEV1 responses to O3 exposure 

and baseline lung function was observed in individuals with asthma, i.e., responses increased 

with severity of disease. Minimal evidence exists suggesting that individuals with asthma have 

smaller O3-induced FEV1 decrements than healthy subjects (3% versus 8%, respectively) 

(Mudway et al., 2001). However, the asthmatics in that study also tended to be older than the 

healthy subjects, which could partially explain their lesser response since FEV1 responses to O3

exposure diminish with age. Individuals with asthma also had significantly more neutrophils in 

the BALF (18 hours postexposure) than similarly exposed healthy individuals (Peden et al., 

1997; Scannell et al., 1996; Basha et al., 1994). Furthermore, a study examining the effects of O3

on individuals with atopic asthma and healthy controls reported that greater numbers of 

neutrophils, higher levels of cytokines and hyaluronan, and greater expression of macrophage 

cell-surface markers were observed in induced sputum of atopic asthmatics compared with 

healthy controls (Hernandez et al., 2010). Differences in O3-induced epithelial cytokine 

expression were noted in bronchial biopsy samples from asthmatics and healthy controls (Bosson 

et al., 2003). Cell-surface marker and cytokine expression results, and the presence of 

hyaluronan, are consistent with O3 having greater effects on innate and adaptive immunity in 

these asthmatic individuals. In addition, studies have demonstrated that O3 exposure leads to 

increased bronchial reactivity to inhaled allergens in mild allergic asthmatics (Kehrl et al., 1999; 

Jorres et al., 1996) and to the influx of eosinophils in individuals with pre-existing allergic 

disease (Vagaggini et al., 2002; Peden et al., 1995). Taken together, these results point to several 

mechanistic pathways which could account for the enhanced sensitivity to O3 in subjects with 

asthma (see Section 5.4.2.2 in the ISA). 

Toxicological studies provide additional evidence of the biological basis for the greater 

effects of O3 among those with asthma or AHR (U.S. EPA, 2013, section 8.2.2). In animal 

toxicological studies, an asthmatic phenotype is modeled by allergic sensitization of the 

respiratory tract. Many of the studies that provide evidence that O3 exposure is an inducer of 

AHR and remodeling utilize these types of animal models. For example, a series of experiments 

in infant rhesus monkeys have shown these effects, but only in monkeys sensitized to house dust 

mite allergen. Similarly, adverse changes in pulmonary function were demonstrated in mice 

exposed to O3; enhanced inflammatory responses were in rats exposed to O3, but only in animals 

sensitized to allergen. In general, it is the combined effects of O3 and allergic sensitization which 

result in measurable effects on pulmonary function. In a pulmonary fibrosis model, exposure O3

for 5 days increased pulmonary inflammation and fibrosis, along with the frequency of 
3-81 

bronchopneumonia in rats. Thus, short-term exposure to O3 may enhance damage in a previously 

injured lung (U.S. EPA, 2013, section 8.2.2). 

In the 2006 O3 AQCD, the potential for individuals with asthma to have greater risk of 

O3-related health effects was supported by a number of controlled human exposure studies, 

evidence from toxicological studies, and a limited number of epidemiologic studies. In section 

8.2.2, the ISA reports that in the recent epidemiologic literature some, but not all, studies report 

greater risk of health effects among individuals with asthma. Studies examining effect measure 

modification of the relationship between short-term O3 exposure and altered lung function by 

corticosteroid use provided limited evidence of O3-related health effects. However, recent studies 

of behavioral responses have found that studies do not take into account individual behavioral 

adaptations to forecasted air pollution levels (such as avoidance and reduced time outdoors), 

which may underestimate the observed associations in studies that examined the effect of O3

exposure on respiratory health (Neidell and Kinney, 2010). This could explain some 

inconsistency observed among recent epidemiologic studies. The evidence from controlled 

human exposure studies provides support for increased detriments in FEV1 and greater 

inflammatory responses to O3 in individuals with asthma than in healthy individuals without a 

history of asthma. The collective evidence for increased risk of O3-related health effects among 

individuals with asthma from controlled human exposure studies is supported by recent 

toxicological studies which provide biological plausibility for heightened risk of asthmatics to 

respiratory effects due to O3 exposure. Overall, the ISA finds there is adequate evidence for 

asthmatics to be an at-risk population. 

Children 

Children are considered to be at greater risk from O3 exposure because their respiratory 

systems undergo lung growth until about 18-20 years of age and are therefore thought to be 

intrinsically more at risk for O3-induced damage (U.S. EPA, 2006). It is generally recognized 

that children spend more time outdoors than adults, and therefore would be expected to have 

higher exposure to O3 than adults. The ventilation rates also vary between children and adults, 

particularly during moderate/heavy activity. Children aged 11 years and older and adults have 

higher absolute ventilation rates than children aged 1-11 years. However, children have higher 

ventilation rates relative to their lung volumes, which tends to increase dose normalized to lung 

surface area. Exercise intensity has a substantial effect on ventilation rate, with high intensity 

activities resulting in nearly double the ventilation rate during moderate activity among children 

and those adults less than 31 years of age. For more information on time spent outdoors and 

ventilation rate differences by age group, see Section 4.4.1 in the ISA (U.S. EPA, 2013). 

The 1996 O3 AQCD reported clinical evidence that children, adolescents, and young 

adults (<18 years of age) appear, on average, to have nearly equivalent spirometric responses to 
3-82 

O3 exposure, but have greater responses than middle-aged and older adults (U.S. EPA, 1996). 

Symptomatic responses (e.g., cough, shortness of breath, pain on deep inspiration) to O3

exposure, however, appear to increase with age until early adulthood and then gradually decrease 

with increasing age (U.S. EPA, 1996). Complete lung growth and development is not achieved 

until 18-20 years of age in women and the early 20s for men; pulmonary function is at its 

maximum during this time as well. 

Recent epidemiologic studies have examined different age groups and their risk to 

O3-related respiratory hospital admissions and emergency department (ED) visits. Evidence for 

greater risk in children was reported in several studies. A study in Cyprus of short-term O3

concentrations and respiratory hospital admissions (HA) detected possible effect measure 

modification by age with a larger association among individuals < 15 years of age compared 

with those > 15 years of age; the effect was apparent only with a 2-day lag (Middleton et al., 

2008). Similarly, a Canadian study of asthma-ED visits reported the strongest O3-related 

associations among 5- to 14-year olds compared to the other age groups (ages examined 0-75+) 

(Villeneuve et al., 2007). Greater O3-associated risk in asthma-related ED visits were also 

reported among children (<15 years) as compared to adults (15 to 64 years) in a study from 

Finland (Halonen et al., 2009). A study of New York City hospital admissions demonstrated an 

increase in the association between O3 exposure and asthma-related hospital admissions for 6- to 

18-year olds compared to those < 6 years old and those > 18 years old (Silverman and Ito, 2010). 

When examining long-term O3 exposure and asthma HA among children, associations were 

determined to be larger among children 1 to 2 years old compared to children 2 to 6 years old 

(Lin et al., 2008b). A few studies reported positive associations among both children and adults 

and no modification of the effect by age. 

The evidence reported in epidemiologic studies is supported by recent toxicological 

studies which observed O3-induced health effects in immature animals. Early life exposures of 

multiple species of laboratory animals, including infant monkeys, resulted in changes in 

conducting airways at the cellular, functional, ultra-structural, and morphological levels. The 

studies conducted on infant monkeys are most relevant for assessing effects in children. Carey et 

al. (2007) conducted a study of O3 exposure in infant rhesus macaques, whose respiratory tract 

closely resemble that of humans. Monkeys were exposed either acutely or in episodes designed 

to mimic human exposure. All monkeys acutely exposed to O3 had moderate to marked 

necrotizing rhinitis, with focal regions of epithelial exfoliation, numerous infiltrating neutrophils, 

and some eosinophils. The distribution, character, and severity of lesions in episodically exposed 

infant monkeys were similar to that of acutely exposed animals. Neither exposure protocol for 

the infant monkeys produced mucous cell metaplasia proximal to the lesions, an adaptation 

observed in adult monkeys exposed in another study (Harkema et al., 1987). Functional and 
3-83 

cellular changes in conducting airways were common manifestations of exposure to O3 among 

both the adult and infant monkeys (Plopper et al., 2007). In addition, the lung structure of the 

conducting airways in the infant monkeys was significantly stunted by O3 and this aberrant 

development was persistent 6 months postexposure (Fanucchi et al., 2006). 

Age may also affect the inflammatory response to O3 exposure. Toxicological studies 

reported that the difference in effects among younger lifestage test animals may be due to 

age-related changes in antioxidants levels and sensitivity to oxidative stress. Further discussion 

of these studies may be found in section 8.3.1.1 of the ISA (U.S. EPA, 2013, p. 8-18). 

The previous and recent human clinical and toxicological studies reported evidence of 

increased risk from O3 exposure for younger ages, which provides coherence and biological 

plausibility for the findings from epidemiologic studies. Although there was some inconsistency, 

generally, the epidemiologic studies reported positive associations among both children and 

adults or just among children. The interpretation of these studies is limited by the lack of 

consistency in comparison age groups and outcomes examined. However, overall, the 

epidemiologic, controlled human exposure, and toxicological studies provide adequate evidence 

that children are potentially at increased risk of O3-related health effects. 

Older Adults 

The ISA notes that older adults are at greater risk of health effects associated with O3

exposure through a variety of intrinsic pathways (U.S. EPA, 2013, section 8.3.1.2). In addition, 

older adults may differ in their exposure and internal dose. Older adults were outdoors for a 

slightly longer proportion of the day than adults aged 18-64 years. Older adults also have 

somewhat lower ventilation rates than adults aged 31 - less than 61 years. For more information 

on time spent outdoors and ventilation rate differences by age group, see Section 4.4 in the ISA 

(U.S. EPA, 2013). The gradual decline in physiological processes that occur with aging may lead 

to increased risk of O3-related health effects (U.S. EPA, 2006). Respiratory symptom responses 

to O3 exposure appears to increase with age until early adulthood and then gradually decrease 

with increasing age (U.S. EPA, 1996); lung function responses to O3 exposure also decline from 

early adulthood (U.S. EPA, 1996). The reductions of these responses with age may put older 

adults at increased risk for continued O3 exposure. In addition, older adults, in general, have a 

higher prevalence of preexisting diseases compared to younger age groups and this may also lead 

to increased risk of O3-related health effects (U.S. EPA, 2013, section 8.3.1.2). With the number 

of older Americans increasing in upcoming years (estimated to increase from 12.4% of the U.S. 

population to 19.7% between 2000 to 2030, which is approximately 35 million and 71.5 million 

individuals, respectively) this group represents a large population potentially at risk of O3-related 

health effects (SSDAN CensusScope, 2010; DeNavas-Walt et al., 2011). 
3-84 

The majority of recent studies reported greater effects of short-term O3 exposure and 

mortality among older adults, which is consistent with the findings of the 2006 O3 AQCD. A 

study (Medina-Ramón and Schwartz, 2008) conducted in 48 cities across the U.S. reported larger 

effects among adults ≥ 65 years old compared to those < 65 years; further investigation of this 

study population revealed a trend of O3-related mortality risk that gets larger with increasing age 

starting at age 51 (Zanobetti and Schwartz, 2008a). Another study conducted in 7 urban centers 

in Chile reported similar results, with greater effects in adults ≥ 65 years old (Cakmak et al., 

2007). More recently, a study conducted in the same area reported similar associations between 

O3 exposure and mortality in adults aged < 64 years old and 65 to 74 years old, but the risk was 

increased among older age groups (Cakmak et al., 2011). A study performed in China reported 

greater effects in populations ≥ 45 years old (compared to 5 to 44 year olds), with statistically 

significant effects present only among those ≥ 65 years old (Kan et al., 2008). An Italian study 

reported higher risk of all-cause mortality associated with increased O3 concentrations among 

individuals ≥85 year old as compared to those 35 to 84 years old (Stafoggia et al., 2010). The Air 

Pollution and Health: A European and North American Approach (APHENA) project examined 

the association between O3 exposure and mortality for those <75 and ≥ 75 years of age. In 

Canada, the associations for all-cause and cardiovascular mortality were greater among those 

≥75 years old. In the U.S., the association for all-cause mortality was slightly greater for those 

<75 years of age compared to those ≥75 years old in summer-only analyses. No consistent 

pattern was observed for CVD mortality. In Europe, slightly larger associations for all-cause 

mortality were observed in those <75 years old in all-year and summer-only analyses. Larger 

associations were reported among those <75 years for CVD mortality in all-year analyses, but 

the reverse was true for summer-only analyses (Katsouyanni et al., 2009). 

With respect to epidemiologic studies of O3 exposure and hospital admissions, a positive 

association was reported between short-term O3 exposure and respiratory hospital admissions for 

adults ≥ 65 years old but not for those adults aged 15 to 64 years (Halonen et al., 2009). In the 

same study, no association was observed between O3 concentration and respiratory mortality 

among those ≥ 65 years old or those 15 to 64 years old. No modification by age (40 to 64 year 

olds versus > 64 year olds) was observed in a study from Brazil examining O3 levels and COPD 

ED visits. 

Although some outcomes reported mixed findings regarding an increase in risk for older 

adults, recent epidemiologic studies report consistent positive associations between short-term 

O3 exposure and mortality in older adults. The evidence from mortality studies is consistent with 

the results reported in the 2006 O3 AQCD and is supported by toxicological studies providing 

biological plausibility for increased risk of effects in older adults. Also, older adults may be 

experiencing increased exposure compared to younger adults. Overall, the ISA (U.S. EPA, 2013) 
3-85 

concludes adequate evidence is available indicating that older adults are at increased risk of 

O3-related health effects. 

People with Diets Lower in Vitamins C and E 

Diet was not examined as a factor potentially affecting risk in previous O3 AQCDs, but 

recent studies have examined modification of the association between O3 and health effects by 

dietary factors. Because O3 mediates some of its toxic effects through oxidative stress, the 

antioxidant status of an individual is an important factor that may contribute to increased risk of 

O3-related health effects. Supplementation with vitamins C and E has been investigated in a 

number of studies as a means of inhibiting O3-mediated damage. 

Two epidemiologic studies have examined effect measure modification by diet and found 

evidence that certain dietary components are related to the effect O3 has on respiratory outcomes. 

In one recent study the effects of fruit/vegetable intake and Mediterranean diet were examined. 

Increases in these food patterns, which have been noted for their high vitamins C and E and 

omega-3 fatty acid content, were positively related to lung function in asthmatic children living 

in Mexico City, and modified by O3 exposure (Romieu et al., 2009). Another study examined 

supplementation of the diets of asthmatic children in Mexico with vitamins C and E (SienraMonge et al., 2004). Associations were detected between short-term O3 exposure and nasal 

airway inflammation among children in the placebo group but not in those receiving the 

supplementation. 

The epidemiologic evidence is supported by controlled human exposure studies, 

discussed in section 8.4.1 of the ISA (U.S. EPA, 2013), that have shown that the first line of 

defense against oxidative stress is antioxidants-rich extracellular lining fluid (ELF) which 

scavenge free radicals and limit lipid peroxidation. Exposure to O3 depletes antioxidant levels in 

nasal ELF probably due to scrubbing of O3; however, the concentration and the activity of 

antioxidant enzymes either in ELF or plasma do not appear to be related to O3 responsiveness. 

Controlled studies of dietary antioxidant supplementation have demonstrated some protective 

effects of α-tocopherol (a form of vitamin E) and ascorbate (vitamin C) on spirometric measures 

of lung function after O3 exposure but not on the intensity of subjective symptoms and 

inflammatory responses. Dietary antioxidants have also afforded partial protection to asthmatics 

by attenuating postexposure bronchial hyperresponsiveness. Toxicological studies discussed in 

section 8.4.1 of the ISA (U.S. EPA, 2013) provide evidence of biological plausibility to the 

epidemiologic and controlled human exposure studies. 

There is adequate evidence that individuals with diets lower in vitamins C and E are at 

risk for O3-related health effects. The evidence from epidemiologic studies is supported by 

controlled human exposure and toxicological studies. 
3-86 

Outdoor Workers 

Studies included in the 2006 O3 AQCD reported that individuals who participate in 

outdoor activities or work outside to be a population at increased risk based on consistently 

reported associations between O3 exposure and respiratory health outcomes in these groups (U.S. 

EPA, 2006). Outdoor workers are exposed to ambient O3 concentrations for a greater period of 

time than individuals who spend their days indoors. As discussed in Section 4.7 of the ISA (U.S. 

EPA, 2013) outdoor workers sampled during the work shift had a higher ratio of personal 

exposure to fixed-site monitor concentrations than health clinic workers who spent most of their 

time indoors. Additionally, an increase in dose to the lower airways is possible during outdoor 

exercise due to both increases in the amount of air breathed (i.e., minute ventilation) and a shift 

from nasal to oronasal breathing. The association between FEV1 responses to O3 exposure and 

minute ventilation is discussed more fully in Section 6.2.3.1 of the 2006 O3 AQCD. 

Previous studies have shown that increased exposure to O3 due to outdoor work leads to 

increased risk of O3-related health effects, specifically decrements in lung function (U.S. EPA, 

2006). The strong evidence from the 2006 O3 AQCD which demonstrated increased exposure, 

dose, and ultimately risk of O3-related health effects in this population supports the conclusion 

that there is adequate evidence to indicate that increased exposure to O3 through outdoor work 

increases the risk of O3-related health effects. 

In some cases, it is difficult to determine a factor that results in increased risk of effects. 

For example, previous assessments have included controlled human exposure studies in which 

some healthy individuals demonstrate greater O3-related health effects compared to other healthy 

individuals. Intersubject variability has been observed for lung function decrements, 

symptomatic responses, pulmonary inflammation, AHR, and altered epithelial permeability in 

healthy adults exposed to O3 and these results tend to be reproducible within a given individual 

over a period of several months indicating differences in the intrinsic responsiveness. In many 

cases the reasons for the variability is not clear. This may be because one or some of the factors 

described above have not been evaluated in studies, or it may be that additional, unidentified 

factors influence individual responses to O3 (U.S. EPA, 2013, section 8.5). 

As discussed in chapter 8 of the ISA the challenges and limitations in evaluating the 

factors that can increase risk for experiencing O3-related health effects may contribute to a lack 

of information about the factors that may increase risk from O3 exposures. This lack of 

information may contribute to conclusions that evidence for some factors, such as sex, SES, and 

obesity provided “suggestive” evidence of increased risk, or that for a number of factors the 

evidence was inadequate to draw conclusions about potential increase in risk of effects. Overall, 

the factors for which the ISA concludes there is adequate evidence of increased risk for 
3-87 

experiencing O3-related effects were related to asthma, lifestage (children and older adults), 

genetic variability, dietary factors, and working outdoors. 

3.1.5.2 Size of At-Risk Populations and Lifestages in the United States 

One consideration in the assessment of potential public health impacts is the size of 

various population groups for which there is adequate evidence of increased risk for health 

effects associated with O3-related air pollution exposure. The factors for which the ISA judged 

the evidence to be “adequate” with respect to contributing to increased risk of O3-related effects 

among various populations and lifestages included: asthma; childhood and older adulthood; diets 

lower in vitamins C and E; certain genetic variants and, working outdoors (U.S. EPA, 2013, 

section 8.5). 

With regard to asthma, Table 3-7 below summarizes information on the prevalence of 

current asthma by age in the U.S. adult population in 2010 (Schiller et al., 2012; children - 

Bloom et al., 2011). Individuals with current asthma constitute a fairly large proportion of the 

population, including more than 25 million people. Asthma prevalence tends to be higher in 

children than adults. 

Within the U.S., approximately 8.2% of adults have reported currently having asthma 

(Schiller et al., 2012) and 9.5% of children have reported currently having asthma (Bloom et al., 

2011). Table 3-12 below provides more detailed information on prevalence of asthma by age in 

the U.S. 
3-88 

Table 3-7. Prevalence of asthma by age in the U.S. 

Age (years) N (in thousands) Percent 

0-4 1,285 6.0 

5-11 3,020 10.5 

12-17 2,672 10.9 

18-44 8,902 8.1 

45-64 6,704 8.4 

65-74 1,849 8.7 

75+ 1,279 7.4 

Asthma prevalence is reported for “still has asthma” 

Source: Statistics for adults: Schiller et al. (2012); Statistics for children: Bloom et al. (2011) 

With regard to lifestages, based on U.S. census data from 2010 (Howden and Meyer, 

2011), about 74 million people, or 24% of the U.S. population, are under 18 years of age and 

more than 40 million people, or about 13% of the U.S. population, are 65 years of age or older. 

Hence, a large proportion of the U.S. population, more than 33%, is included in age groups that 

are considered likely to be at increased risk for health effects from ambient O3 exposure. 

With regard to dietary factors, no statistics are available to estimate the size of an at-risk 

population based on nutritional status. 

With regard to outdoor workers, in 2010 approximately 11.7% of the total number of 

people (143 million people) employed, or about 16.8 million people, worked outdoors one or 

more day per week (based on worker surveys).63 Of these approximately 7.4% of the workforce, 

or about 7.8 million people, worked outdoors three or more days per week. 

The health statistics data illustrate what is known as the “pyramid” of effects. At the top 

of the pyramid, there are approximately 2.5 million deaths from all causes per year in the U.S. 

population, with about 250 thousand respiratory-related deaths (CDC-WONDER64). For 

respiratory health diseases, there are nearly 3.3 million hospital discharges per year (HCUP65), 

 

63 The O*NET program is the nation's primary source of occupational information. Central to the project is the 

O*NET database, containing information on hundreds of standardized and occupation-specific descriptors. The 

database, which is available to the public at no cost, is continually updated by surveying a broad range of workers 

from each occupation. http://www.onetcenter.org/overview.html 

http://www.onetonline.org/find/descriptor/browse/Work_Context/4.C.2/ 

64 http://wonder.cdc.gov/ 

65 http://www.hcup-us.ahrq.gov/ 
3-89 

8.7 million respiratory ED visits (HCUP, 2007), 112 million ambulatory 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. 

3.1.5.3 Averting Behavior 

The activity pattern of individuals is an important determinant of their exposure (ISA, 

U.S. EPA, 2013, section 4.4.1). Variation in O3 concentrations among various 

microenvironments means that the amount of time spent in each location, as well as the level of 

activity, will influence an individual’s exposure to ambient O3. Activity patterns vary both 

among and within individuals, resulting in corresponding variations in exposure across a 

population and over time. Individuals can reduce their exposure to O3 by altering their behaviors, 

such as by staying indoors, being active outdoors when air quality is better, and by reducing their 

activity levels or reducing the time being active outdoors on high-O3 days (U.S. EPA, 2013, 

section 4.4.2). The evidence in this topic area, while not addressed in the 2006 AQCD, is 

evaluated in the ISA for this review. 

The widely reported Air Quality Index (AQI) conveys advice to the public, and 

particularly at-risk populations, on reducing exposure on days when ambient levels of common 

air pollutants are elevated (www.airnow.gov). The AQI describes the potential for health effects 

from O3 (and other individual pollutants) in six color-coded categories of air-quality, ranging 

from Good (green), Moderate (yellow), Unhealthy for Sensitive Groups (orange), Unhealthy 

(red), and Very Unhealthy (purple), and Hazardous (maroon). Levels in the unhealthy ranges 

(i.e., Unhealthy for Sensitive Groups and above) come with recommendations about reducing 

exposure. Forecasted and actual AQI values for O3 are reported to the public during the O3

season. The AQI advisories explicitly state that children, older adults, people with lung disease, 

and people who are active outdoors, may be at greater risk from exposure to O3. People are 

advised to reduce exposure depending on the predicted O3 levels and the likelihood of risk. This 

advice includes being active outdoors when air quality is better, and reducing activity levels or 

reducing the time being active outdoors on high-O3 days. Staying indoors to reduce exposure is 

not recommended until air quality reaches the Very Unhealthy or Hazardous categories. 

Evidence of individual averting behaviors in response to AQI advisories has been found 

in several studies, including activity pattern and epidemiologic studies, especially for the at-risk 

populations, such as children, older adults, and people with asthma, who are targeted by the 

advisories. Such effects are less pronounced in the general population, possibly due to the 

opportunity cost of behavior modification. Epidemiologic evidence from a study (Neidell and 
3-90 

Kinney, 2010) conducted in the 1990’s in Los Angeles, CA reports increased asthma hospital 

admissions among children and older adults when O3 alert days (1-hour max O3 concentration 

>200 ppb) were excluded from the analysis of daily hospital admissions and O3 concentrations 

(presumably thereby eliminating averting behavior based on high O3 forecasts). The lower rate of 

admissions observed when alert days were included in the analysis suggests that estimates of 

health effects based on concentration-response functions that do not account for averting 

behavior may be biased towards the null (U.S. EPA, 2013, section 4.4.2). 

3.2 AIR QUALITY-, EXPOSURE-, AND RISK-BASED CONSIDERATIONS 

In order to inform judgments about the public health impacts of O3-related health effects, 

the HREA has developed and applied models to estimate human exposures to O3 and O3-

associated health risks across the United States, with a specific focus on urban case study areas 

(U.S. EPA, 2014).66 The HREA uses photochemical modeling to adjust air quality from the 

2006-2010 O3 seasons to just meet the current and alternative standards for the 2006-2008 and 

2008-2010 periods.67 In this section, staff considers estimates of short-term O3 exposures and 

estimates of health risks associated with short- and long-term O3 exposures, for air quality 

adjusted to just meet the current O3 standard. In section 3.2.1, we consider the implications for 

exposure and risk estimates of the approach used in the HREA to adjust air quality. Sections 

3.2.2 and 3.2.3 discuss our exposure-based and risk-based considerations, respectively. In these 

sections we specifically consider the following question: 

 What are the nature and magnitude of O3 exposures and health risks remaining 

upon adjusting recent air quality to just meet the current O3 standard, and what are 

the important uncertainties associated with those exposure and risk estimates? 

3.2.1 Consideration of the Adjusted Air Quality Used in Exposure and Risk 

Assessments 

In the first draft HREA for this review, as in the last review, the EPA relied upon 

quadratic rollback to adjust hourly O3 concentrations in urban case study areas to just meet the 

current O3 standard (U.S. EPA, 2014). Although the quadratic rollback method reproduces 

 

66 The 15 urban case study areas analyzed for exposures are Atlanta, Baltimore, Boston, Chicago, Cleveland, Dallas, 

Denver, Detroit, Houston, Los Angeles, New York, Philadelphia, Sacramento, St. Louis, and Washington, DC. 

Morbidity and mortality risk estimates are presented for these same areas, with the exception of Chicago, Dallas, 

and Washington, DC. The HREA also presents a national scale mortality risk assessment for unadjusted (recent) air 

quality. This national-scale assessment, which focuses on existing air quality conditions and does not estimate the 

health risks associated with just meeting the current or alternative standards, can provide perspective on the 

relationship between national-scale O3 public health impacts and impacts estimated in specific urban areas. 

67 Three-year periods are used recognizing that the current standard is the average across three years of the annual 

fourth-highest daily maximum 8-hour average concentration. 
3-91 

historical patterns of air quality changes better than some alternative methods, it relies on 

statistical relationships without explicitly accounting for atmospheric chemistry and precursor 

emissions (U.S. EPA, 2014, Chapter 4). An important drawback of the quadratic rollback 

approach, recognized in the first draft HREA (U.S. EPA, 2012b), is that it forces all monitors in 

an assessment area to exhibit the same response when air quality is adjusted. It does not allow for 

the spatial or temporal heterogeneity in responses that result from the non-linear atmospheric 

chemistry that influences ambient O3 concentrations (U.S. EPA, 2014, Chapter 4). Because 

quadratic rollback does not account for physical and chemical atmospheric processes, or the 

sources of emissions precursors that lead to O3 formation, a backstop or “floor” must be used 

when applying quadratic rollback to just meet current or alternative standards to ensure that 

estimated O3 is not reduced in a manner inconsistent with O3 chemistry, such as to reduce 

concentrations below that associated with background sources (U.S. EPA, 2014, Chapter 4). 

Consistent with recommendations from the National Research Council of the National 

Academies (NRC, 2008), the HREA uses a photochemical model to estimate sensitivities of O3

to changes in precursor emissions, in order to estimate ambient O3 concentrations that would just 

meet the current and alternative standards (U.S. EPA, 2014, Chapter 4).68 For the urban case 

study areas evaluated in the HREA, this model-based adjustment approach was set up to estimate 

hourly O3 concentrations at each monitor location when modeled U.S. anthropogenic precursor 

emissions (i.e., NOX, VOC)69 were reduced to estimate air quality that just meets the current and 

alternative O3 standards.70 

As discussed in Chapter 4 of the HREA (U.S. EPA, 2014), this approach models the 

physical and chemical atmospheric processes that influence ambient O3 concentrations. 

Compared to the quadratic rollback approach, it provides more realistic estimates of the spatial 

and temporal responses of O3 to reductions in precursor emissions. These improved estimates 

avoid many of the limitations inherent in the quadratic rollback method, including the 

requirement that all monitors in an assessment area exhibit the same response upon air quality 

 

68 The HREA uses the CMAQ photochemical model instrumented with the higher order direct decoupled method 

(HDDM) to estimate ozone concentrations that would occur with the achievement of the current and alternative O3

standards (U.S. EPA, 2014, Chapter 4). 

69 Exposure and risk analyses for most urban case study areas focus on reducing NOX emissions alone (NOX

emissions were reduced by about 10 to 85% for the current standard, and up to about 95% for alternatives). In most 

of the urban case study areas, reducing VOC emissions did not alter the NOX emissions reductions required to just 

meet the current or alternative standards. However, in Chicago and Denver, reductions in VOC emissions allowed 

for smaller NOX emissions reductions. Therefore, exposure and risk analyses for Chicago and Denver are based on 

reductions in emissions of both NOX and VOC (U.S. EPA, 2014, section 4.3.3.1, Table 4-3). 

70 Although this chapter focuses on the current standard, our overarching considerations regarding adjusted air 

quality also apply to alternative standards simulated in the HREA. Alternative standards are discussed in chapter 4 

of this PA. 
3-92 

adjustment to the current and/or alternative standards. Because adjusted air quality scenarios are 

based on reducing only U.S. anthropogenic emissions, this approach also does not require the 

specification of background concentrations as a rollback “floor” (U.S. EPA, 2014, section 4.3.3). 

The use of this model-based air quality adjustment approach in the HREA has important 

implications for the patterns of ambient O3 concentrations estimated in urban case study areas. 

Specifically, in locations and time periods when NOX is predominantly contributing to O3

formation (e.g., downwind of important NOX sources, where the highest O3 concentrations often 

occur), model-based adjustment to the current and alternative standards decreases estimated 

ambient O3 concentrations compared to recent monitored concentrations (U.S. EPA, 2014, 

section 4.3.3.2). In contrast, in locations and time periods when NOX is predominantly 

contributing to O3 titration (e.g., in urban centers with high concentrations of NOX emissions, 

where ambient O3 concentrations are often suppressed and thus relatively low71), model-based 

adjustment increases ambient O3 concentrations compared to recent measured concentrations 

(U.S. EPA, 2014, section 4.3.3.2) (Chapter 2, above). 

Within urban case study areas, the overall impacts of model-based air quality adjustment 

are to reduce relatively high ambient O3 concentrations (i.e., concentrations at the upper ends of 

ambient distributions) and to increase relatively low O3 concentrations (i.e., concentrations at the 

lower ends of ambient distributions) (U.S. EPA, 2014, section 4.3.3.2, Figures 4-9 and 4-10). 

Seasonal means of daily concentrations generally exhibit only modest changes upon air quality 

adjustment, reflecting the seasonal balance between daily decreases and increases in ambient 

concentrations (U.S. EPA, 2014, Figures 4-9 and 4-10). The resulting compression in 

distributions of ambient O3 concentrations is evident in all of the urban case study areas 

evaluated, though the degree of compression varies considerably across areas (U.S. EPA, 2014, 

Figures 4-9 and 4-10). 

Adjusted patterns of O3 air quality have important implications for exposure and risk 

estimates in urban case study areas. Estimates influenced largely by the upper ends of the 

distribution of ambient concentrations (i.e., exposures of concern and lung function risk 

estimates, as discussed in sections 3.2.2 and 3.2.3.1 below) will decrease with model-adjustment 

to the current and alternative standards. In contrast, seasonal risk estimates influenced by the full 

distribution of ambient O3 concentrations (i.e., epidemiology-based risk estimates, as discussed 

in section 3.2.3.2 below) either increase or decrease in response to air quality adjustment, 

 

71 Titration is also prominent during time periods when photochemistry is limited, such as at night and on cool, 

cloudy days. 
3-93 

depending on the balance between the daily decreases in high O3 concentrations and increases in 

low O3 concentrations.72 

We further consider the implications of the spatial and temporal patterns of adjusted air 

quality within the context of exposure (section 3.2.2) and risk (section 3.2.3) estimates for O3

concentrations adjusted to just meet the current standard. As discussed below (section 3.2.3.2), 

these altered patterns are particularly important to consider when interpreting epidemiologybased risk estimates. 

3.2.2 Exposure-Based Considerations 

The exposure assessment presented in the HREA (U.S. EPA, 2014, Chapter 5) provides 

estimates of the number of people exposed to various concentrations of ambient O3, while at 

specified exertion levels. The HREA estimates exposures in 15 urban case study areas for 

school-age children (ages 5 to 18), asthmatic school-age children, asthmatic adults, and older 

adults, reflecting the strong evidence indicating that these populations are potentially at increased 

risk for O3-attributable effects (EPA, 2013, Chapter 8; section 3.1.2, above). An important 

purpose of these exposure estimates is to provide perspective on the extent to which air quality 

adjusted to just meet the current O3 NAAQS could be associated with exposures to O3

concentrations reported to result in respiratory effects.73 Estimates of such “exposures of 

concern” provide perspective on the potential public health impacts of O3-related effects, 

including for effects that cannot currently be evaluated in a quantitative risk assessment (e.g., 

airway inflammation). 

In the absence of large scale exposure studies that encompass the general population, as 

well as at-risk populations, modeling is the preferred approach to estimating exposures to O3. 

The use of exposure modeling also facilitates the estimation of exposures resulting from ambient 

air concentrations differing from those in exposure studies (e.g., concentrations just meeting the 

current standard). In the HREA, population exposures to ambient O3 concentrations are 

estimated using the current version of the Air Pollutants Exposure (APEX) model. The APEX 

model simulates the movement of individuals through time and space and estimates their 

exposures to a given pollutant in indoor, outdoor, and in-vehicle microenvironments (U.S. EPA, 

2014, section 5.1.3). APEX takes into account the most critical factors that contribute to total 

 

72 In addition, because epidemiology-based risk estimates use “area-wide” average O3 concentrations, calculated by 

averaging concentrations across multiple monitors in urban case study areas (section 3.2.3.2 below), risk estimates 

on a given day depend on the daily balance between increasing and decreasing O3 concentrations at individual 

monitors. 

73In addition, the range of modeled personal exposures to ambient O3 provide an essential input to the portion of the 

health risk assessment based on exposure-response functions (for lung function decrements) from controlled human 

exposure studies. The health risk assessment based on exposure-response information is discussed in section 3.2.3, 

below. 
3-94 

human exposure to ambient O3, including the temporal and spatial distributions of people and O3

concentrations throughout an urban area, the variation of O3 concentrations within various 

microenvironments, and the effects of exertion on breathing rate in exposed individuals (U.S. 

EPA, 2014, section 5.1.3). To the extent spatial and/or temporal patterns of ambient O3

concentrations are altered upon air quality adjustment, as discussed above, exposure estimates 

reflect population exposures to those altered patterns. 

The HREA estimates 8-hour exposures at or above benchmark concentrations of 60, 70, 

and 80 ppb for individuals engaged in moderate or greater exertion. Benchmarks reflect exposure 

concentrations at which O3-induced respiratory effects are known to occur in some healthy adults 

engaged in moderate, intermittent exertion, based on evidence from controlled human exposure 

studies (section 3.1.2.1 above and U.S. EPA, 2013, section 6.2). The amount of weight to place 

on the estimates of exposures at or above specific benchmark concentrations depends in part on 

the weight of the scientific evidence concerning health effects associated with O3 exposures at 

that concentration. It also depends on judgments about the importance, from a public health 

perspective, of the health effects that are known or can reasonably be inferred to occur as a result 

of exposures at benchmark concentrations (sections 3.1.3, 3.1.5 above). 

As discussed in more detail above (section 3.1.2.1), the health evidence that supports 

evaluating exposures of concern at or above benchmark concentrations of 60, 70, and 80 ppb 

comes from a large body of controlled human exposure studies reporting a variety of respiratory 

effects in healthy adults. The lowest O3 exposure concentration for which controlled human 

exposure studies have reported respiratory effects in healthy adults is 60 ppb, with more 

evidence supporting this benchmark concentration in the current review than in the last review. 

In healthy adults, exposures to 60 ppb O3 have been reported to decrease lung function and to 

increase airway inflammation. Exposures of healthy adults to 70 ppb O3 have been reported to 

result in larger lung function decrements, compared to 60 ppb, as well as in increased respiratory 

symptoms.74 Exposures of healthy adults to 80 ppb O3 have been reported to result in larger lung 

function decrements than following exposures to 60 or 70 ppb, increased airway inflammation, 

increased respiratory symptoms, increased airways responsiveness, and decreased lung host 

defense (section 3.1.2.1, above). As discussed above (section 3.1.3), respiratory effects reported 

following exposures to O3 concentrations of 60, 70, or 80 ppb meet ATS criteria for adverse 

effects, result in effects judged important by CASAC in past reviews, and/or could contribute to 

the clearly adverse effects reported in epidemiologic studies evaluating broader populations. 

Compared to the healthy individuals included in the studies that provided the basis for the 

 

74 As noted above, for the 70 ppb exposure concentration Schelegle et al. (2009) reported that the actual mean 

exposure concentration was 72 ppb. 
3-95 

benchmarks, at-risk populations (e.g., asthmatics, children) are more likely to experience larger 

and/or more serious effects (e.g., U.S. EPA, 2013, p. 6-21). 

In considering estimates of O3 exposures of concern at or above benchmarks of 60, 70, 

and 80 ppb, within the context of the adequacy of the current standard, we first address the 

following specific question: 

 What are the nature and magnitude of the short-term O3 exposures of concern 

remaining upon adjustment of air quality to just meet the current O3 standard? 

In addressing this question, we focus on modeled exposures for school-age children (ages 

5-18) and asthmatic school-age children, two of the at-risk populations identified in the ISA 

(section 3.1.5 above). The percentages of children estimated to experience exposures of concern 

are larger than the percentages estimated for adult populations (i.e., approximately 3-fold larger 

across cities) (U.S. EPA, 2014, sections 5.3.2, 5.3.3 and Figures 5-5 to 5-8). The larger exposure 

estimates for children are due primarily to the larger percentage of children estimated to spend an 

extended period of time being physically active outdoors when O3 concentrations are elevated 

(U.S. EPA, 2014, sections 5.3.2 and 5.4.1). 

Although exposure estimates differ between children and adults, the patterns of results 

across the cities and years are similar among all of the populations evaluated (U.S. EPA, 2014, 

Figures 5-5 to 5-8). Therefore, while we highlight estimates in children, we also note that the 

patterns of exposures estimated for children represent the patterns estimated for adult asthmatics 

and older adults. 

Key results for children are summarized below for air quality adjusted to simulate just 

meeting the current O3 NAAQS (Figures 3-7 to 3-10).75 Estimates for all children and asthmatic 

children are virtually indistinguishable (U.S. EPA, 2014, section 5.3.2). The estimates presented 

in Figures 3-7 to 3-10 below reflect consistent reductions in estimated exposures of concern 

across urban case study areas, relative to recent (i.e., unadjusted) air quality (U.S. EPA, 2014, 

Appendix 5F). When averaged over the years evaluated in the HREA, reductions of up to about 

70% were estimated, compared to recent air quality. These reductions in estimated exposures of 

concern, relative to unadjusted air quality, reflect the consistent reductions in the highest ambient 

O3 concentrations upon air quality adjustment to just meet the current standard (section 3.2.1 

above; U.S. EPA, 2014, Chapter 4). Such reductions in estimated exposures of concern are 

evident throughout urban case study areas, including in urban cores and in surrounding areas 

(U.S. EPA, 2014, section 9.6, Appendix 9A). Figures 3-7 (Average over years) and 3-8 (WorstCase Years) present estimates of one or more exposures of concern. Figures 3-9 (Average over 

years) and 3-10 (Worst-Case Years) present estimates of two or more exposures of concern. 

 

75 Figures 3-7 and 3-8 present estimates of one or more exposures of concern. Figures 3-9 and 3-10 present estimates 

of two or more exposures of concern. 
3-96 

Figure 3-7. Percent of children estimated to experience one or more exposures of concern at or above 60, 

quality adjusted to just meet the current standard - Averaged Over 2006 to 2010. 
3-97 

Figure 3-8. Percent of children estimated to experience one or more exposures of concern at or above 60, 

quality adjusted to just meet the current standard - Worst-Case Year from 2006 to 2010. 
3-98 

Figure 3-9. Percent of children estimated to experience two or more exposures of concern at or above 60, 

quality adjusted to just meet the current standard - Averaged Over 2006 to 2010. 
3-99 

Figure 3-10. Percent of children estimated to experience two or more exposures of concern at or above 60, 

quality adjusted to just meet the current standard - Worst-Case Year from 2006 to 2010. 
3-100 

Based on Figures 3-7 to 3-10 and the associated details described in the HREA (U.S. 

EPA, 2014, Chapter 5), we take note of the following with regard to exposures that are estimated 

to be allowed by the current standard: 

1. For exposures of concern at or above 60 ppb: 

a. On average over the years 2006 to 2010, the current standard is estimated to allow 

approximately 10 to 17% of children in urban case study areas to experience one or 

more exposures of concern at or above 60 ppb. Summing across urban case study 

areas, these percentages correspond to almost 2.5 million children experiencing 

approximately 4 million exposures of concern at or above 60 ppb during a single O3

season. Of these children, almost 250,000 are asthmatics. 

b. On average over the years 2006 to 2010, the current standard is estimated to allow 

approximately 3 to 8% of children in urban case study areas to experience two or 

more exposures of concern to O3 concentrations at or above 60 ppb. Summing across 

the urban case study areas, these percentages correspond to almost 900,000 children 

(including about 90,000 asthmatic children) estimated to experience at least two O3

exposure concentrations at or above 60 ppb during a single O3 season. 

c. In the worst-case years (i.e., those with the largest exposure estimates), the current 

standard is estimated to allow approximately 10 to 26% of children to experience one 

or more exposures of concern at or above 60 ppb, and approximately 4 to 14% to 

experience two or more exposures of concern at or above 60 ppb. 

2. For exposures of concern at or above 70 ppb: 

a. On average over the years 2006 to 2010, the current standard is estimated to allow up 

to approximately 3% of children in urban case study areas to experience one or more 

exposures of concern at or above 70 ppb. Summing across urban case study areas, 

more than 350,000 children (including about 40,000 asthmatic children) are estimated 

to experience O3 exposure concentrations at or above 70 ppb during a single O3

season. 

b. On average over the years 2006 to 2010, the current standard is estimated to allow 

less than 1% of children in urban case study areas to experience two or more 

exposures of concern to O3 concentrations at or above 70 ppb. 

c. In the worst-case years, the current standard is estimated to allow approximately 1 to 

8% of children to experience one or more exposures of concern at or above 70 ppb, 

and up to approximately 2% to experience two or more exposures of concern, at or 

above 70 ppb. 

3. For exposures of concern at or above 80 ppb: The current standard is estimated to allow 

about 1% or fewer children in urban case study areas to experience exposures of concern at 

or above 80 ppb, even in years with the highest exposure estimates. 
3-101 

In further evaluating estimated exposures of concern from the HREA, we next consider 

the following question: 

 What are the important sources of uncertainty associated with exposure estimates? 

Due to variability in responsiveness, only a subset of individuals who experience 

exposures at or above a benchmark concentration can be expected to experience health effects. 

Given the lack of sufficient exposure-response information for most of the health effects that 

informed benchmark concentrations, estimates of the number of people likely to experience 

exposures at or above benchmark concentrations generally cannot be translated into quantitative 

estimates of the number of people likely to experience specific health effects.76 We view healthrelevant exposures as a continuum with greater confidence and certainty about the existence of 

adverse health effects at higher O3 exposure concentrations, and less confidence and greater 

uncertainty as one considers lower exposure concentrations. This view draws from the overall 

body of available health evidence, which indicates that as exposure concentrations increase the 

incidence, magnitude, and severity of effects increases. 

Though we have less confidence in the likelihood of adverse health effects as O3

exposure concentrations decrease, we also note that the controlled human exposure studies that 

provided the basis for health benchmark concentrations have not evaluated at-risk populations. 

Compared to the healthy individuals included in controlled human exposure studies, members of 

at-risk populations (e.g., asthmatics, children) could be more likely to experience adverse effects, 

could experience larger and/or more serious effects, and/or could experience effects following 

exposures to lower O3 concentrations. In considering estimated exposures of concern within the 

context of drawing conclusions on the adequacy of the current standard (section 3.4, below), we 

balance concerns about the potential for adverse health effects, including effects in at-risk 

populations, with our increasing uncertainty regarding the likelihood of such effects following 

exposures to lower O3 concentrations. 

Uncertainties associated with the APEX exposure modeling also have the potential to be 

important in our consideration of the adequacy of the current standard (U.S. EPA, 2014, section 

5.5.2, Table 5-10). For example, the HREA concludes that exposures of concern could be 

underestimated for some individuals who are frequently and routinely active outdoors during the 

warm season (U.S. EPA, section 5.5.2). This could include outdoor workers and children who 

are frequently active outdoors. The HREA specifically notes that long-term diary profiles (i.e., 

monthly, annual) do not exist for such populations, limiting the extent to which APEX outputs 

reflect people who follow similar daily routines resulting in high exposures, over extended 

 

76 The exception to this is lung function decrements, as discussed below (section 3.2.3.1). 
3-102 

periods of time. Thus, exposure estimates generated from the general pool of available diary 

profiles likely do not reflect the most highly exposed individuals in the population. 

In order to evaluate the potential implications of this uncertainty for exposure estimates, 

the HREA reports the results of limited sensitivity analyses using subsets of diaries specifically 

selected to reflect groups spending a larger proportion of time being active outdoors during the 

O3 season. When diaries were selected to mimic exposures that could be experienced by outdoor 

workers, the percent of modeled individuals estimated to experience exposures of concern 

increased compared to other adult populations evaluated. The percent of outdoor workers 

estimated to experience exposures of concern were generally similar to the percentages estimated 

for children (i.e., using the full database of diary profiles) in the worst-case cities and years (i.e., 

cities and years with the highest exposure estimates) (U.S. EPA, 2014, section 5.4.3.2, Figure 5-

14). In addition, when diaries were restricted to children who did not report any time spent inside 

a school or performing paid work (i.e., to mimic children spending particularly large portions of 

their time outdoors during the summer), the number experiencing exposures of concern increased 

by approximately 30% (U.S. EPA, 2014, section 5.4.3.1). Though these sensitivity analyses are 

limited to single urban case study areas, and though there is uncertainty associated with diary 

selection approaches to mimic highly exposed populations, they suggest the possibility that some 

at-risk groups could experience more frequent exposures of concern than indicated by estimates 

based on the full database of activity diary profiles. 

In further considering activity diaries, the HREA also notes growing evidence indicating 

that people can change their behavior in response to high O3 concentrations, reducing the time 

spent being active outdoors (U.S. EPA, 2014, section 5.4.3.3). Commonly termed “averting 

behaviors,” these altered activity patterns could reduce personal exposure concentrations. 

Therefore, the HREA also performed limited sensitivity analyses to evaluate the potential 

implications of averting behavior for estimated exposures of concern. These analyses suggest 

that averting behavior could reduce the percentages of children estimated to experience 

exposures of concern at or above the 60 or 70 ppb benchmark concentrations by approximately 

10 to 30%, with larger reductions possible for the 80 ppb benchmark (U.S. EPA, 2014, Figure 5-

15). As discussed above for other sensitivity analyses, these analyses are limited to a single 

urban case study area and are subject to uncertainties associated with assumptions about the 

prevalence and duration of averting behaviors. However, the results suggest that exposures of 

concern could be overestimated, particularly in children (Neidell et al., 2009; U.S. EPA, 2013, 

Figures 4-7 and 4-8), if the possibility for averting behavior is not incorporated into estimates. 
3-103 

3.2.3 Risk-Based Considerations 

For some health endpoints, there is sufficient scientific evidence and information 

available to support the development of quantitative estimates of O3-related health risks. In the 

last review of the O3 NAAQS, the quantitative health risk assessment estimated O3-related lung 

function decrements, respiratory symptoms, respiratory-related hospital admissions, and nonaccidental and cardiorespiratory-related mortality (U.S. EPA, 2007). In those analyses, both 

controlled human exposure and epidemiologic studies were used for the quantitative assessment 

of O3-related human health risks. 

In the current review, for short-term O3 concentrations the HREA estimates lung function 

decrements; respiratory symptoms in asthmatics; hospital admissions and emergency department 

visits for respiratory causes; and all-cause mortality (U.S. EPA, 2014, Chapters 6 and 7). For 

“long-term” O3 concentrations, the HREA estimates respiratory mortality (U.S. EPA, 2014, 

Chapter 7).77 Estimates of O3-induced lung function decrements are based on exposure modeling, 

as noted above, combined with exposure-response relationships from controlled human exposure 

studies (U.S. EPA, 2014, Chapter 6). Estimates of O3-associated respiratory symptoms; hospital 

admissions and emergency department visits; and mortality are based on concentration-response 

relationships from epidemiologic studies (U.S. EPA, 2014, Chapter 7). As with the exposure 

assessment discussed above, O3-associated health risks are estimated for recent air quality and 

for ambient concentrations adjusted to just meet the current 8-hour O3 NAAQS, based on 2006-

2010 air quality and adjusted precursor emissions. 

Section 3.2.3.1 below discusses risk results for O3-induced lung function decrements 

following short-term exposures, based on exposure modeling results and exposure-response 

relationships from controlled human exposure studies. Section 3.2.3.2 discusses epidemiologybased risk estimates, with a focus on all-cause mortality (short-term O3 concentrations); 

respiratory-related morbidity outcomes (short-term O3 concentrations); and respiratory mortality 

(long-term O3 concentrations). 

3.2.3.1 Risk of Lung Function Decrements 

In the last review, EPA conducted a health risk assessment that produced risk estimates 

for the number and percent of school-aged children, asthmatic school-aged children, and the 

general population experiencing lung function decrements associated with O3 exposures for 12 

urban areas. These estimates were based on exposure-response relationships developed from 

 

77 Risk estimates for “long-term” concentrations are based on the concentration-response relationship identified in 

the study by Jerrett et al. (2009). As discussed above, study authors used April to September averages of 1-hour 

daily maximum O3 concentrations as surrogates for “long-term” exposures. 
3-104 

analysis of data from several controlled human exposure studies, combined with exposure 

estimates developed for children and adults (U.S. EPA, 2007). 

In the current review, the HREA estimates risks of lung function decrements in schoolaged children (ages 5 to 18), asthmatic school-aged children, and the general adult population for 

15 urban case study areas.78 The results presented in the HREA are based on an updated dosethreshold model that estimates FEV1 responses for individuals following short-term exposures to 

O3 (McDonnell et al., 2012), reflecting methodological improvements since the last review (U.S. 

EPA, 2014, section 6.2.4; section 3.1.2.1, above). The impact of the dose threshold is that O3-

induced FEV1 decrements result primarily from exposures on days with ambient O3

concentrations above about 40 ppb (U.S. EPA, 2013, section 6.3.1, Figure 6-9).79

As discussed above (section 3.1.3), for effects such as lung function decrements, which 

are transient and reversible, aspects such as the likelihood that these effects would occur 

repeatedly and would interfere with normal activities are important to consider in making 

judgments about adversity to individuals. As stated in the 2006 Criteria Document (Table 8-3, p. 

8-68), for people with lung disease even moderate functional responses (e.g., FEV1 decrements ≥

10% but < 20%) would likely interfere with normal activities for many individuals, and would 

likely result in more frequent medication use. Moreover, as noted above, in a recent letter to the 

Administrator, the CASAC O3 Panel stated that “'[c]linically relevant effects are decrements > 

10%, a decrease in lung function considered clinically relevant by the American Thoracic 

Society” (Samet, 2011, p.2). The CASAC O3 Panel also stated that: 

[A] 10% decrement in FEV1 can lead to respiratory symptoms, especially in 

individuals with pre-existing pulmonary or cardiac disease. For example, people 

with chronic obstructive pulmonary disease have decreased ventilatory reserve 

(i.e., decreased baseline FEV1) such that a ≥ 10% decrement could lead to 

moderate to severe respiratory symptoms (Samet, 2011, p.7). 

Consistent with this advice from the last review, in the current review CASAC has concluded 

that “estimation of FEV1 decrements of ≥15% is appropriate as a scientifically relevant surrogate 

for adverse health outcomes in active healthy adults, whereas an FEV1 decrement of ≥10% is a 

scientifically relevant surrogate for adverse health outcomes for people with asthma and lung 

disease” (Frey, 2014, p. 3). 

 

78As noted for the exposure assessment above, the 15 cities assessed are Atlanta, Baltimore, Boston, Chicago, 

Cleveland, Dallas, Denver, Detroit, Houston, Los Angeles, New York, Philadelphia, Sacramento, St. Louis, and 

Washington, DC. 

79 Error! Reference source not found. in the HREA shows that more than 90% of daily instances of FEV1

decrements ≥ 10% occur when 8-hr average ambient concentrations are above 40 ppb for 2006 air quality in Los 

Angeles. The distribution of decrements will be different for different study areas, years, and air quality scenarios 

(U.S. EPA, 2014, section 6.3.1). 
3-105 

 In judging the extent to which moderate lung function decrements represent effects that 

should be regarded as adverse to the health status of individuals, in previous NAAQS reviews we 

have also considered the extent to which decrements were experienced repeatedly during the 

course of a year (U.S. EPA, 2007). Although some experts would judge single occurrences of 

moderate responses to be a ‘‘nuisance,’’ especially for healthy individuals,80 a more general 

consensus view of the adversity of such moderate responses emerges as the frequency of 

occurrence increases. Thus in the past EPA has judged that repeated occurrences of moderate 

responses, even in otherwise healthy individuals, may be considered to be adverse since they 

could well set the stage for more serious illness (61 FR 65723). The CASAC panel in the 1997 

NAAQS review expressed a consensus view that these “criteria for the determination of an 

adverse physiological response were reasonable” (Wolff, 1995). 

 The HREA estimates risks of moderate to large lung function decrements, defined as 

FEV1 decrements > 10%, > 15%, or > 20%. In evaluating these lung function risk estimates 

within the context of considering the adequacy of the current O3 standard, we first consider the 

following specific question: 

 What are the nature and magnitude of lung function risks remaining upon just meeting 

the current O3 standard? 

In considering risks of O3-induced FEV1 decrements, we focus on the percent of children 

estimated to experience decrements > 10, 15, and 20%, noting that the percentage of asthmatic 

children estimated to experience such decrements is virtually indistinguishable from the 

percentage estimated for all children.81 Compared to children, smaller percentages of adults are 

estimated to experience O3-induced FEV1 decrements (U.S. EPA, 2014, section 6.3.1, Table 6-

4). As for exposures of concern (see above), the patterns of results across urban case study areas 

and over the years evaluated are similar in children and adults (U.S. EPA, 2014, Appendix 6E). 

Therefore, while we highlight estimates in children, we note that these results are also 

representative of the patterns estimated for adult populations. 

Key results for children are summarized below for air quality adjusted to just meet the 

current O3 NAAQS (Figures 3-11 to 3-14).82 The estimates presented in Figures 3-11 to 3-14 

below reflect consistent reductions across urban case study areas in the percent of children 

estimated to experience O3-induced lung function decrements, relative to recent (i.e., unadjusted) 

air quality (U.S. EPA, 2014, Appendix 6B). When averaged over the years evaluated in the 

 

80Though not all experts, as indicated by the advice received on this issue from past CASAC O3 Panels (Samet, 

2011). 

81 Though see below for discussion of uncertainty in lung function responses of children and asthmatics. 

82 Figures 3-11 and 3-12 present estimates of one or more decrements. Figures 3-13 and 3-14 present estimates of 

two or more decrements. 
3-106 

HREA, reductions of up to about 40% were estimated compared to recent air quality. These 

reductions reflect the consistent decreases in relatively high ambient O3 concentrations upon 

adjustment to just meet the current standard (section 3.2.1 above; U.S. EPA, 2014, Chapter 4).83

Such reductions in estimated lung function risks are evident throughout urban case study areas, 

including in urban cores and in surrounding areas (U.S. EPA, 2014, section 9.6). Figures 3-11 

(Average over years) and 3-12 (Worst-Case Years) present estimates of one or more O3-induced 

lung function decrements. Figures 3-13 (Average over years) and 3-14 (Worst-Case Years) 

present estimates of two or more decrements. 

 

83 As noted above, the impact of the dose threshold is that O3-induced FEV1 decrements result primarily from days 

with average ambient O3 concentrations above about 40 ppb. 
3-107 

Figure 3-11. Percent of school-aged children (5-18 yrs) estimated to experience one or more days with FEV

15, or 20% with air quality adjusted to just meet the current standard – Averaged over 2006 
3-108 

Figure 3-12. Percent of school-aged children (5-18 yrs) estimated to experience one or more days with FEV

15, or 20% with air quality adjusted to just meet the current standard – Worst-Case Year fro
3-109 

Figure 3-13. Percent of school-aged children (aged 5-18 yrs) estimated to experience two or more days with 

10, 15, or 20% with air quality adjusted to just meet the current standard – Averaged over 20
3-110 

Figure 3-14. Percent of school-aged children (5-18 yrs) estimated to experience two or more days with FEV

15, or 20% with air quality adjusted to just meet the current standard - Worst-Case Year fro
3-111 

Based on Figures 3-11 to 3-14 and the associated details described in the HREA (U.S. 

EPA, 2014, Chapter 6), we take note of the following with regard to lung function decrements 

estimated to be allowed by the current standard: 

1. For FEV1 decrements > 10%: 

a. On average over the years 2006 to 2010, the current standard is estimated to allow 

approximately 14 to 19% of children in urban case study areas to experience one or 

more lung function decrements > 10%. Summing across urban case study areas, this 

corresponds to approximately 3 million children experiencing 15 million O3-induced 

lung function decrements > 10% during a single O3 season. Of these children, about 

300,000 are asthmatics. 

b. On average over the years 2006 to 2010, the current standard is estimated to allow 

approximately 8 to 12% of children in urban case study areas to experience two or 

more O3-induced lung function decrements > 10%. Summing across the urban case 

study areas, this corresponds to almost 2 million children (including almost 200,000 

asthmatic children) estimated to experience two or more O3-induced lung function 

decrements greater than 10% during a single O3 season. 

c. In the worst-case years, the current standard is estimated to allow approximately 17 to 

22% of children in urban case study areas to experience one or more lung function 

decrements > 10%, and approximately 10 to 14% to experience two or more O3-

induced lung function decrements > 10%. 

2. For FEV1 decrements > 15%: 

a. On average over the years 2006 to 2010, the current standard is estimated to allow 

approximately 3 to 5% of children in urban case study areas to experience one or 

more lung function decrements > 15%. Summing across urban case study areas, this 

corresponds to over 750,000 children (including approximately 80,000 asthmatic 

children) estimated to experience at least one O3-induced lung function decrement > 

15% during a single O3 season. 

b. On average over the years 2006 to 2010, the current standard is estimated to allow 

approximately 2 to 3% of children in urban case study areas to experience two or 

more O3-induced lung function decrements > 15%. 

c. In the worst-case years, the current standard is estimated to allow approximately 4 to 

7% of children in urban case study areas to experience one or more lung function 

decrements > 15%, and approximately 2 to 4% to experience two or more O3-induced 

lung function decrements > 15%. 

3. For FEV1 decrements > 20%: 

a. On average over the years 2006 to 2010, the current standard is estimated to allow 

approximately 1 to 2% of children in urban case study areas to experience one or 

more lung function decrements > 20%. Summing across urban case study areas, this 

corresponds to almost 300,000 children (including approximately 30,000 asthmatic 
3-112 

children) estimated to experience at least one O3-induced lung function decrement > 

20% during a single O3 season. 

b. On average over the years 2006 to 2010, the current standard is estimated to allow 

less than about 1% of children in urban case study areas to experience two or more 

O3-induced lung function decrements > 20%. 

c. In the worst-case years, the current standard is estimated to allow approximately 2 to 

3% of children to experience one or more lung function decrements > 20%, and less 

than 2% to experience two or more O3-induced lung function decrements > 20%. 

In further considering estimated lung function risks from the HREA, we next consider the 

following question: 

 What are the important sources of uncertainty associated with lung function risk 

estimates? 

In addition to the uncertainties noted above for exposure estimates, the HREA identifies 

several key uncertainties associated with estimates of O3-induced lung function decrements. An 

uncertainty with particular potential to impact our consideration of risk estimates in this Policy 

Assessment stems from the lack of exposure-response information in children. In the absence of 

controlled human exposure data for children, risk estimates are based on the assumption that 

children exhibit the same lung function response following O3 exposures as healthy 18 year olds 

(i.e., the youngest age for which controlled human exposure data is available) (U.S. EPA, 2014, 

section 6.5.3). This assumption was justified in part by the findings of McDonnell et al. (1985), 

who reported that children 8-11 year old experienced FEV1 responses similar to those observed 

in adults 18-35 years old. In addition, as discussed in the ISA (U.S. EPA, 2013, section 6.2.1), 

summer camp studies of school-aged children reported O3-induced lung function decrements 

similar in magnitude to those observed in controlled human exposure studies using adults. In 

extending the risk model to children, the HREA fixes the age term in the model at its highest 

value, the value for age 18. This approach could result in either over- or underestimates of O3-

induced lung function decrements in children, depending on how children compare to the adults 

used in controlled human exposure studies (U.S. EPA, 2014, section 6.5.3). 

A related source of uncertainty is that the risk assessment estimates O3-induced 

decrements in asthmatics using the exposure-response relationship developed from data collected 

from healthy individuals. Although the evidence has been mixed (U.S. EPA, 2013, section 

6.2.1.1), several studies have reported larger O3-induced lung function decrements in asthmatics 

than in non-asthmatics (Kreit et al., 1989; Horstman et al., 1995; Jorres et al., 1996; Alexis et al., 

2000). Consistent with the findings of the ISA (U.S. EPA, 2013, section 6.2.1.1), CASAC noted 

that “[a]sthmatic subjects appear to be at least as sensitive, if not more sensitive, than nonasthmatic subjects in manifesting ozone-induced pulmonary function decrements” (Frey, 2014, 
3-113 

p. 4). To the extent asthmatics experience larger O3-induced lung function decrements than the 

healthy adults used to develop exposure-response relationships, the HREA could underestimate 

the impacts of O3 exposures on lung function in asthmatics, including asthmatic children. The 

HREA notes that the magnitude this uncertainty might have on risk estimates remains unknown 

at this time (U.S. EPA, 2014, section 6.5.4). 

3.2.3.2 Estimated Health Risks Associated with Short- or Long-Term O3 Exposures, 

Based on Epidemiologic Studies 

Risk estimates based on epidemiologic studies can provide perspective on the most 

serious O3-associated public health outcomes (e.g., mortality, hospital admissions, emergency 

department visits) in populations that often include at-risk groups. The HREA estimates O3-

associated risks in 12 urban case study areas84 using concentration-response relationships drawn 

from epidemiologic studies. These concentration-response relationships are based on “areawide” average O3 concentrations.85 The HREA estimates risks for the years 2007 and 2009 in 

order to provide estimates of risk for a year with generally higher O3 concentrations (2007) and a 

year with generally lower O3 concentrations (2009) (U.S. EPA, 2014, section 7.1.1). 

In the last review, epidemiologic-based risks were estimated for O3 concentrations above 

mean “policy-relevant background concentrations.” As discussed above (Chapter 2), policyrelevant background (PRB) concentrations were defined as the distribution of ozone 

concentrations that would be observed in the U.S. in the absence of anthropogenic (man-made) 

emissions of ozone precursor emissions (e.g., VOC, CO, NOx) in the U.S., Canada, and Mexico. 

This approach provided a focus on O3 concentrations “that can be controlled by U.S. regulations 

(or through international agreements with neighboring countries)” (U.S. EPA, 2007, pp. 2-48 to 

2-54). 

As in the last review, we recognize that ambient O3 concentrations, and therefore O3-

associated health risks, result from precursor emissions from various types of sources. Based on 

the air quality modeling discussed above in chapter 2, approximately 30 to 60% of average 

daytime O3 during the warm season (i.e., daily maximum 8-hour concentrations averaged from 

April to October) is attributable to precursor emissions from U.S. anthropogenic sources (section 

2.4.4). The remainder is attributable to precursor emissions from international anthropogenic 

 

84 The 12 urban areas evaluated are Atlanta, Baltimore, Boston, Cleveland, Denver, Detroit, Houston, Los Angeles, 

New York, Philadelphia, Sacramento, and St. Louis. 

85 In the epidemiologic studies that provide the health basis for HREA risk assessments, concentration-response 

relationships are based on daytime O3 concentrations, averaged across multiple monitors within study areas. These 

daily averages are used as surrogates for the spatial and temporal patterns of exposures in study populations. 

Consistent with this approach, the HREA epidemiologic-based risk estimates also utilize daytime O3 concentrations, 

averaged across monitors, as surrogates for population exposures. In this PA, we refer to these averaged 

concentrations as “area-wide” O3 concentrations. Area-wide concentrations are discussed in more detail in section 

3.1.4, above. 
3-114 

sources and natural sources. Because the HREA characterizes health risks from all O3, regardless 

of source, risk estimates reflect emissions from U.S. anthropogenic, international anthropogenic, 

and natural sources. 

In evaluating epidemiology-based risk estimates within the context of the adequacy of the 

current standard, we first consider the following question: 

 What are the nature and magnitude of the O3-associated mortality and morbidity risks 

remaining upon adjustment of air quality to just meet the current O3 standard? 

In addressing this question, we note that the HREA estimates mortality and morbidity 

risks associated with just meeting the current standard by applying concentration-response 

relationships from epidemiologic studies to the entire distributions of adjusted “area-wide” 

average O3 concentrations present in urban case study areas (U.S. EPA, 2014, Chapter 7). 

Implicit in this approach to estimating risks is the assumption that concentration-response 

relationships are linear over those distributions. Therefore, as noted in section 3.2.1, when air 

quality is adjusted to just meet the current standard, risk estimates are influenced by the 

decreases in area-wide O3 concentrations at the upper ends of warm season distributions and the 

increases in area-wide O3 concentrations at the lower ends of those distributions (U.S. EPA, 

2014, section 4.3.3.2, Figures 4-9 and 4-10).86 When the decreases and increases are of the same 

magnitude, they result in the same degree of change in estimated risks, though opposite in 

direction. Therefore, seasonal estimates of O3-associated mortality and morbidity risks either 

increase or decrease in response to air quality adjustment, depending on the seasonal balance 

between the modeled daily decreases in high area-wide O3 concentrations and increases in low 

area-wide O3 concentrations. One consequence is that the estimated impacts on mortality and 

morbidity risks of adjusting air quality to just meet the current standard are more modest, and 

less directionally consistent across urban case study areas, than on either exposures of concern or 

O3-induced lung function decrements. 

In the remainder of this section, we consider estimates of total (non-accidental) mortality 

and respiratory morbidity associated with short-term O3 concentrations, and respiratory mortality 

associated with “long-term” O3 concentrations. 

Total Mortality – Short-Term O3

Risk estimates for total mortality are based on concentration-response relationships 

described by Smith et al. (2009). To generate risk estimates, the HREA uses “area-wide” 

 

86 On a given day, area-wide O3 concentrations and estimated risks decrease when the sum of the changes at 

monitors with decreasing O3 (e.g., downwind of important NOX sources, where the highest O3 concentrations often 

occur) are larger than the sum of the changes at monitors with increasing O3 (e.g., often in urban centers with high 

concentrations of NOX emissions, where ambient O3 concentrations are suppressed and thus relatively low). Areawide O3 concentrations and estimated risks increase when the opposite occurs. 
3-115 

averages of daily maximum 8-hour O3 concentrations over the full monitoring periods in urban 

case study areas. When 2007 air quality was adjusted to the current standard (the year with 

generally “higher” O3-associated risks), 10 of 12 urban case study areas exhibited either 

decreases or virtually no change in estimates of the number of O3-associated deaths (U.S. EPA, 

2014, Appendix 7B). Increases were estimated in two of the urban case study areas (Houston, 

Los Angeles) (U.S. EPA, 2014, Appendix 7B).87 

Figure 3-15 below presents estimates of O3-associated all-cause mortality in urban case 

study areas for 2007 and 2009, with air quality adjusted to just meet the current O3 standard. The 

HREA estimates that upon just meeting the current standard, O3 could be associated with from 

0.8 to 4.1% of all-cause mortality across the urban case study areas. This corresponds to 

approximately 60 to 3,200 O3-associated deaths per season in individual urban case study areas, 

and approximately 7,000 to 7,500 O3-associated deaths per season summed over the 12 urban 

case study areas (U.S. EPA, 2014, Tables 7-7 and 7-8). 

Figure 3-15. Percent of all-cause mortality associated with O3 for air quality adjusted to 

just meet the current standard. 

In considering the risk estimates presented in Figure 3-15, which are based on applying 

linear concentration-response relationships to the full distributions of daily 8-hour “area-wide” 

 

87 For 2009 (i.e., the year with generally lower O3 concentrations), changes in risk were generally smaller than in 

2007 (i.e., most changes about 2% or smaller). Increases were estimated for Houston, Los Angeles, and New York 

City. 
3-116 

O3 concentrations, we note the ISA conclusion that there is less certainty in the shape of 

concentration-response functions for area-wide O3 concentrations at the lower ends of warm 

season distributions (i.e., below about 20 to 40 ppb depending on the O3 metric, health endpoint, 

and study population) (U.S. EPA, 2013, section 2.5.4.4). We also recognize that for the range of 

health endpoints evaluated, controlled human exposure and animal toxicological studies provide 

greater certainty in the increased incidence, magnitude, and severity of effects at higher exposure 

concentrations (discussed in sections 3.1.2.2 and 3.1.4.2, above).88 Thus, in addition to 

considering estimates of total O3-associated risks, we also consider the extent to which risks are 

associated with days with higher, versus lower, area-wide O3 concentrations. 

Figure 3-16 presents risk estimates, summed across urban case study areas, for days with 

area-wide concentrations at or above 20, 40, and 60 ppb. Daytime O3 concentrations in the upper 

portion of the distribution of area-wide concentrations (e.g., at or above 40 or 60 ppb) are 

estimated to be associated with hundreds to thousands of deaths per year in urban case study 

areas.89 

Figure 3-16. Estimated O3-associated deaths attributable to various area-wide average O3

concentrations, with air quality adjusted to just meet current standard. 

 

88 As discussed in section 3.1.4.2, as ambient concentrations increase the potential for exposures to higher O3

concentrations also increases. Thus with increasing ambient concentrations, controlled human exposure and animal 

toxicological studies support the increased incidence, magnitude, and severity of O3-attributable effects. 

89 The relatively small proportion of O3-associated deaths attributable to days with area-wide concentrations of 60 

ppb or greater reflects the relatively small proportion of days with such elevated area-wide concentrations. 
3-117 

Respiratory Mortality – “Long-Term” O3 

The HREA estimates the risk of respiratory mortality associated with long-term O3

exposures, based on the study by Jerrett et al. (2009) (U.S. EPA, 2014, Chapter 7). As discussed 

above (section 3.1.4.3), Jerrett et al. (2009) reported that when seasonal averages of 1-hour daily 

maximum O3 concentrations ranged from 33 to 104 ppb, there was no statistically significant 

deviation from a linear concentration-response relationship between O3 and respiratory mortality 

across 96 U.S. cities (U.S. EPA, 2013, section 7.7). However, the authors reported “limited 

evidence” for an effect threshold at an O3 concentration of 56 ppb (p=0.06). In communications 

with EPA staff (described in Sasser, 2014), the study authors indicated that it is not clear whether 

a threshold model is a better predictor of respiratory mortality than the linear model, and that 

“considerable caution should be exercised in accepting any specific threshold.” Consistent with 

this communication, the HREA estimated respiratory mortality associated with long-term O3

concentrations based on the linear model from the published study, and in a series of sensitivity 

analyses with models that included thresholds ranging from 40 to 60 ppb (U.S. EPA, 2014, 

Figure 7-9). 

To generate risk estimates, the HREA uses “area-wide” averages of 1-hour daily 

maximum O3 concentrations during the warm season (April to September). When 2007 air 

quality was adjusted to just meet the current standard (i.e., the year with generally higher O3

concentrations) all 12 of the urban case study areas exhibited decreases in estimated O3-

associated respiratory mortality (i.e., compared to recent, unadjusted air quality). For 2009 

adjusted air quality (i.e., the year with generally lower O3 concentrations), urban case study areas 

exhibited either no change in estimated risk, or decreases in risk that were smaller than those for 

2007 (U.S. EPA, 2014, Appendix 7B, Tables 7B-6 and 7B-7). Risk estimates based on the linear 

model, for air quality adjusted to just meet the current standard, are presented below in Figure 3-

17. 
3-118 

Figure 3-17. Percent of baseline respiratory mortality estimated to be associated with longterm O3. 

Based on a concentration-response function that is linear over the entire distribution of 

long-term O3 concentrations, O3 is estimated to be associated with approximately 16 to 21% of 

respiratory deaths in urban case study areas during the warm season. This corresponds to 

approximately 300 to 2,100 O3-associated deaths per season in individual urban case study areas, 

and a total of approximately 8,000 to 9,000 O3-associated deaths summed across all 12 case 

study areas. Based on threshold models, HREA sensitivity analyses indicate that the number of 

respiratory deaths associated with long-term O3 concentrations could potentially be considerably 

lower (i.e., by more than 75% if a threshold exists at 40 ppb, and by about 98% if a threshold 

exists at 56 ppb) (U.S. EPA, 2014, Figure 7-9). 

Hospital Admissions, Emergency Department Visits, and Asthma Exacerbations 

Risk estimates for respiratory-related hospital admissions, emergency department visits, 

and asthma exacerbations associated with air quality adjusted to just meet the current standard 

are based on several studies, as presented in Table 7-2 of the HREA (U.S. EPA, 2014).90

Estimates indicate that O3-associated respiratory-related hospital admissions generally account 

for approximately 2 to 3% of total respiratory-related admissions in urban case study locations. 

Depending on the city, this corresponds to 10’s to 100’s of O3-associated hospital admissions per 

season. Estimates indicate that O3-associated respiratory-related emergency department visits 

 

90As with respiratory mortality above, the HREA does not characterize distributions of respiratory morbidity risks 

over distributions of ambient O3 concentrations. Therefore, in considering respiratory morbidity risks we evaluate 

estimates of total risk. 
3-119 

account for approximately 3 to 20% of total respiratory-related emergency department visits in 

Atlanta and New York City (corresponding to thousands of visits per season in these two cities), 

and that O3-associated asthma exacerbations account for approximately 15 to 30% of total 

exacerbations in Boston (30,000 to 80,000 exacerbations per season). Full estimates are 

presented in Tables 7-9 to 7-11 in the HREA (U.S. EPA, 2014). 

Based on the detailed information presented in Chapter 7 of the HREA (U.S. EPA, 2014), 

we note the following key observations: 

1. In focusing on total risk, the current standard is estimated to allow thousands of O3-

associated deaths per year in the urban case study areas. These estimates are based on 

concentration-response functions from epidemiologic studies that used either 8-hour daily 

O3 concentrations (total mortality associated with short-term O3) or seasonal averages of 

1-hour daily O3 concentrations (respiratory mortality associated with long-term O3). 

2. In focusing on the risks associated with the upper portions of distributions of ambient 

concentrations, the current standard is estimated to allow hundreds to thousands of O3-

associated deaths per year in the urban case study areas. These estimates are based on 

concentration-response functions from an epidemiologic study that evaluated associations 

between 8-hour daily O3 concentrations and total mortality. 

3. In urban case study areas, the current standard is estimated to allow tens to thousands of 

O3-associated morbidity events per year. Distributions of O3-associated morbidity over 

distributions of ambient O3 concentrations would likely be similar to mortality, though 

the HREA did not analyze such distributions for morbidity endpoints. 

In further considering estimated O3-associated mortality and morbidity risks from the 

HREA, we next consider the following question: 

 What are the important sources of uncertainty associated with mortality and morbidity 

risk estimates? 

Compared to estimates of O3 exposures of concern and estimates of O3-induced lung 

function decrements (discussed above), the HREA conclusions reflect somewhat lower 

confidence in epidemiologic-based risk estimates, given important uncertainties. In particular, 

the HREA highlights the unexplained heterogeneity in effect estimates between locations, the 

potential for exposure measurement errors, and uncertainty in the interpretation of the shape of 

concentration-response functions at lower O3 concentrations (U.S. EPA, 2014, section 9.6). The 

HREA also concludes that lower confidence should be placed in the results of the assessment of 

respiratory mortality risks associated with long-term O3 exposures, primarily because that 

analysis is based on only one study (even though that study is well-designed) and because of the 

uncertainty in that study about the existence and level of a potential threshold in the 

concentration-response function (U.S. EPA, 2014, section 9.6). This section discusses some of 

the key uncertainties in epidemiologic-based risk estimates, with a focus on uncertainties that can 
3-120 

have particularly important implications for our consideration of epidemiology-based risk 

estimates in this PA 

Estimating air quality that just meets the current standard based on modeled responses to 

reductions in NOX emissions generally reduces O3-associated mortality and morbidity risks in 

locations and time periods with relatively high ambient O3 concentrations and increases risks in 

locations and time periods with relatively low concentrations. When evaluating uncertainties in 

epidemiologic risk estimates, it is important to consider the extent to which the pattern of air 

quality changes in response to reductions in NOX emissions is representative of trends in ambient 

O3; the extent to which estimated changes in risks in urban case study areas are representative of 

the changes that would be experienced broadly across the U.S. population; and the extent to 

which the O3 response to reductions in precursor emissions could differ with emissions reduction 

strategies that are different from those used in REA risk estimates. 

To evaluate the first issue, the HREA conducted a national analysis evaluating trends in 

monitored ambient O3 concentrations during a time period when the U.S. experienced large-scale 

reductions in NOX emissions (i.e., 2001 to 2010). Analyses of trends in monitored O3 indicate 

that over such a time period, the upper end of the distribution of monitored O3 concentrations 

(i.e., indicated by the 95th percentile) generally decreased in urban and non-urban locations 

across the U.S. (U.S. EPA, 2014, Figure 8-29). During this same time period, median O3

concentrations decreased in suburban and rural locations, and in some urban locations. However, 

median concentrations increased in some large urban centers (U.S. EPA, 2014, Figure 8-28). As 

discussed in the REA, and above (II.C.1), these increases in median concentrations likely reflect 

the increases in relatively low O3 concentrations that can occur near important sources of NOX

upon reductions in NOX emissions (U.S. EPA, 2014, section 8.2.3.1). These patterns of 

monitored O3 during a period when the U.S. experienced large reductions in NOX emissions are 

qualitatively consistent with the modeled responses of O3 to reductions in NOX emissions. 

To evaluate the second issue, the HREA conducted national air quality modeling 

analyses. These analyses estimated the proportion of the U.S. population living in locations 

where seasonal averages of daily O3 concentrations are estimated to decrease in response to 

reductions in NOX emissions, and the proportion living in locations where such seasonal 

averages are estimated to increase. Given the strong relationship between changes in seasonal 

averages of daily O3 concentrations and changes in seasonal risk estimates, this analysis informs 

consideration of the extent to which the risk results in urban case study areas represent the U.S. 

population as a whole. This representativeness analysis indicates that the 12 urban case study 

areas may not represent the response of O3 in other populated areas of the U.S., including 

suburban areas, smaller urban areas, and rural areas. This analysis also indicates that the majority 

of the U.S. population lives in locations where reducing NOX emissions would be expected to 
3-121 

result in decreases in warm season averages of daily maximum 8-hour ambient O3

concentrations. One implication of these results is that HREA risk estimates for the urban case 

study areas may understate the average reduction in O3-associated mortality and morbidity risk 

that would be experienced across the U.S. population upon reducing NOX emissions (U.S. EPA, 

2014, sections 8.2.3.2 and 8.4). 

To evaluate the third issue, the HREA assesses the O3 air quality response to reducing 

both NOX and VOC (i.e., in addition to assessing reductions in NOX emissions alone) for a 

subset of seven urban case study areas. As noted above (section 3.2.1), in most of these urban 

case study areas the inclusion of VOC emissions reductions did not alter the NOX emissions 

reductions required to meet the current or alternative standards.91 However, the addition of VOC 

reductions generally resulted in larger decreases in mid-range O3 concentrations (25th to 75th

percentiles) (U.S. EPA, 2014, Appendix 4D, section 4D-4.7).92 In addition, in all seven of the 

urban case study areas evaluated, the increases in low O3 concentrations were smaller for the 

NOX/VOC scenarios than the NOX alone scenarios (U.S. EPA, 2014, Appendix 4D, section 4D4.7). This was most apparent for Denver, Houston, Los Angeles, New York, and Philadelphia. 

Given the impacts on total risk estimates of increases in low O3 concentrations, these results 

suggest that in some locations better-optimized emissions reduction strategies could result in 

larger reductions in O3-associated mortality and morbidity than indicated in the HREA core 

estimates. 

Section 7.4 of the HREA also highlights some additional uncertainties associated with 

epidemiologic-based risk estimates (U.S. EPA, 2014). This section of the HREA identifies and 

discusses sources of uncertainty and presents a qualitative evaluation of key parameters that can 

introduce uncertainty into risk estimates (U.S. EPA, 2014, Table 7-4). For several of these 

parameters the HREA also presents quantitative sensitivity analyses (U.S. EPA, 2014, section 

7.5.3). Of the uncertainties discussed in Chapter 7 of the HREA, those related to the application 

of concentration-response functions from epidemiologic studies can have particularly important 

implications for our consideration of epidemiology-based risk estimates in this PA, as discussed 

below. 

An important uncertainty is the shape of concentration-response functions at low ambient 

O3 concentrations (U.S. EPA, 2014, Table 7-4).93 Consistent with the ISA conclusion that there 

 

91 The exception is Chicago and Denver, for which the HREA risk estimates are based on reductions in both NOX

and VOC (U.S. EPA, 2014, section 4.3.3.1). Emissions of NOX and VOC were reduced by equal percentages, a 

scenario not likely to reflect the optimal combination for reducing risks. 92 This was the case for all of the urban case study areas evaluated, with the exception of New York (U.S. EPA, 

2014, Appendix 4D, section 4D-4.7). 

93 A related uncertainty is the existence, or not, of a threshold. The HREA addresses this issue for long-term O3 by 

evaluating risks in models that include potential thresholds (see above). 
3-122 

is no discernible population threshold in O3-associated health effects, the HREA estimates 

epidemiology-based mortality and morbidity risks for entire distributions of ambient O3

concentrations, with the assumption that concentration-response relationships remain linear over 

those distributions. In addition, in recognition of the ISA conclusion that certainty in the shape of 

O3 concentration-response functions decreases at low ambient concentrations, the HREA also 

estimates distributions of total mortality incidence for various portions of the distribution of 

ambient O3 concentrations. In this PA, we consider both types of risk estimates while 

recognizing that we have greater certainty in the increased incidence and severity of O3-

attributable effects at higher ambient O3 concentrations (which drive higher exposure 

concentrations, section 3.2.2 above), as compared to lower concentrations.94 

The HREA also notes important uncertainties associated with using a concentrationresponse relationship developed for a particular population in a particular location to estimate 

health risks in different populations and locations (U.S. EPA, 2014, Table 7-4). As discussed 

above, concentration-response relationships derived from epidemiologic studies reflect the 

spatial and temporal patterns of population exposures during the study. The HREA applies 

concentration-response relationships from epidemiologic studies to adjusted air quality in study 

areas that are different from, and often larger in spatial extent than, the areas used to generate the 

relationships. This approach ensures the inclusion of the actual non-attainment monitors that 

often determine the magnitude of emissions reductions for the air quality adjustments throughout 

the urban case study areas. This approach also allows the HREA to estimate patterns of health 

risks more broadly across a larger area, including a broader range of air quality concentrations 

and a larger population. The HREA notes that it is not possible to quantify the impacts of this 

uncertainty on risk estimates in most urban case study locations, though the HREA notes that 

mortality effect estimates for different portions of the New York City CBSA-based assessment 

area vary by a factor of almost 10 (U.S. EPA, 2014, section 7.5.3). 

An additional, related uncertainty is that associated with applying concentration-response 

functions from epidemiologic studies to adjusted air quality. Concentration-response functions 

from the O3 epidemiologic studies used in the HREA are based on associations between day to 

day variation in “area-wide” O3 concentrations (i.e., averaged across multiple monitors) and 

variation in health effects. Epidemiologic studies use these area-wide O3 concentrations, which 

reflect the particular spatial and temporal patterns of ambient O3 present in study locations, as 

surrogates for the pattern of O3 exposures experienced by study populations. To the extent 

adjusting O3 concentrations to just meet the current standard results in important alterations in 

 

94 As discussed above, we also consider the potential implications of the existence of a threshold in the association 

between long-term O3 and respiratory mortality. 
3-123 

the spatial and/or temporal patterns of ambient O3, there is uncertainty in the appropriateness of 

applying concentration-response functions from epidemiologic studies to estimate health risks 

associated with adjusted O3 air quality.95 In particular, this uncertainty could be important to the 

extent that (1) factors associated with space modify the effects of O3 on health or (2) spatial 

mobility is a key driver of individual-level exposures. Although the impact of this uncertainty on 

risk estimates cannot be quantified (U.S. EPA, 2014, Table 7-4), it has the potential to become 

more important as air quality adjustment results in larger changes in spatial and temporal patterns 

of ambient O3 concentrations across urban case study areas. 

There is also uncertainty related specifically to the public health importance of the 

increases in relatively low O3 concentrations following air quality adjustment. This uncertainty 

relates to the fact that HREA risk estimates are equally influenced by decreasing high 

concentrations and increasing low concentrations, when the increases and decreases are of equal 

magnitude. Even on days with increases in relatively low area-wide average concentrations, 

resulting in increases in estimated risks, some portions of the urban case study areas could 

experience decreases in high O3 concentrations. To the extent it is reasonable to conclude that 

O3-attributable effects are more strongly supported for higher ambient concentrations (see 

above), likely resulting in higher exposure concentrations for some portions of study areas, the 

impacts on risk estimates of increasing low O3 concentrations reflect an important source of 

uncertainty. 

The use of a national concentration-response function to estimate respiratory mortality 

associated with long-term O3 is a source of uncertainty. Risk estimates generated in sensitivity 

analyses using region-specific effect estimates differ substantially from the core estimates based 

on a single national-level effect estimate (U.S. EPA, 2014; Table 7-14). Furthermore, the risk 

estimates generated using the regional effect estimates display considerable variability across 

urban case study areas (U.S. EPA, 2014; Table 7-14), reflecting the substantial variability in the 

underlying effect estimates (see Jerrett et al., 2009, Table 4). While the results of the HREA 

sensitivity analyses evaluating this uncertainty point to the potential for regional heterogeneity in 

the long-term risk estimates, the relatively large confidence intervals associated with regional 

effect estimates resulted in the HREA conclusion that staff does not have confidence in the 

regionally-based risk estimates themselves. 

Finally, we note the HREA does not quantify any reductions in risk that could be 

associated with reductions in the ambient concentrations of pollutants other than O3, resulting 

from control of NOX. For example as discussed in chapter 2 of this PA, NOx emissions 

contribute to ambient NO2, and NOx and VOCs can contribute to secondary formation of PM2.5 

 

95 As discussed above (section 3.2.1), decreasing modeled NOX emissions to just meet the current standard can 

dramatically alter the spatial and temporal patterns of ambient O3 concentrations across urban case study areas. 
3-124 

constituents, including ammonium sulfate (NH4SO4), ammonium nitrate (NH4NO3), and organic 

carbon (OC). Therefore, at some times and in some locations, control strategies that would 

reduce NOX emissions (i.e., to meet an O3 standard) could reduce ambient concentrations of NO2

and PM2.5, resulting in health benefits beyond those directly associated with reducing ambient O3

concentrations.96 

3.3 CASAC ADVICE AND PUBLIC COMMENTERS’ VIEWS ON THE 

ADEQUACY OF THE CURRENT STANDARD 

Beyond the evidence- and risk/exposure-based information discussed above, staff has 

also taken into account the comments and advice of CASAC, based on their reviews of the ISA, 

the HREA and PA, as well as comments provided by public commenters. The range of views 

summarized here generally reflects differing judgments as to the relative weight to place on 

various types of evidence, the exposure- and risk-based information, and the associated 

uncertainties, as well as differing judgments about the importance of various O3-related health 

effects from a public health perspective. 

Following the 2008 decision to revise the primary O3 standard by setting the level at 

0.075 ppm (75 ppb), CASAC strongly questioned whether the standard met the requirements of 

the CAA, further described below. In September 2009, EPA announced its intention to 

reconsider the 2008 standards, issuing a notice of proposed rulemaking in January 2010 (FR 75 

2938). Soon after, EPA solicited CASAC review of that proposed rule and in January 2011 

solicited additional advice. This proposal was based on the scientific and technical record from 

the 2008 rulemaking, including public comments and CASAC advice and recommendations. As 

further described in section 1.2.2 above, EPA in the fall of 2011 did not revise the standard as 

part of the reconsideration process but decided to coordinate further proceedings on the 

reconsideration rulemaking with this ongoing periodic review. Accordingly, in this section we 

describe CASAC’s advice related to the 2008 final decision and the subsequent reconsideration, 

as well as its advice on the NAAQS review that was initiated in September 2008. 

In April 2008, the members of the CASAC Ozone Review Panel sent a letter to EPA 

stating “[I]n our most-recent letters to you on this subject—dated October 2006 and March 

2007—the CASAC unanimously recommended selection of an 8-hour average Ozone NAAQS 

within the range of 0.060 to 0.070 parts per million [60 to 70 ppb] for the primary (human 

health-based) Ozone NAAQS” (Henderson, 2008). The letter continued: 

 

96 We expect little focus by states on controlling NOX for purposes of controlling PM2.5 given the more efficient 

control of PM2.5 through reduction of SO2 and direct PM2.5 emissions in most locations. Thus, consideration in this 

review of reductions in ambient PM2.5 resulting from putative NOX control would not double-count PM2.5 emission 

reductions. 
3-125 

The CASAC now wishes to convey, by means of this letter, its additional, 

unsolicited advice with regard to the primary and secondary Ozone NAAQS. In 

doing so, the participating members of the CASAC Ozone Review Panel are 

unanimous in strongly urging you or your successor as EPA Administrator to 

ensure that these recommendations be considered during the next review cycle for 

the Ozone NAAQS that will begin next year ... numerous medical organizations 

and public health groups have also expressed their support of these CASAC 

recommendations’ ... [The CASAC did] not endorse the new primary ozone 

standard as being sufficiently protective of public health. The CASAC—as the 

Agency’s statutorily-established science advisory committee for advising you on 

the national ambient air quality standards—unanimously recommended 

decreasing the primary standard to within the range of 0.060–0.070 ppm [60 to 

70 ppb]. It is the Committee’s consensus scientific opinion that your decision to 

set the primary ozone standard above this range fails to satisfy the explicit 

stipulations of the Clean Air Act that you ensure an adequate margin of safety for 

all individuals, including sensitive populations. 

In response to EPA’s solicitation of their advice on the Agency’s proposed rulemaking as 

part of the reconsideration, CASAC conveyed support (Samet, 2011). 

 CASAC fully supports EPA’s proposed range of 0.060 – 0.070 parts per million 

(ppm) for the 8-hour primary ozone standard. CASAC considers this range to be 

justified by the scientific evidence as presented in the Air Quality Criteria for 

Ozone and Related Photochemical Oxidants (March 2006) and Review of the 

National Ambient Air Quality Standards for Ozone: Policy Assessment of 

Scientific and Technical Information, OAQPS Staff Paper (July 2007). As stated 

in our letters of October 24, 2006, March 26, 2007 and April 7, 2008 to former 

Administrator Stephen L. Johnson, CASAC unanimously recommended selection 

of an 8-hour average ozone NAAQS within the range proposed by EPA (0.060 to 

0.070 ppm). In proposing this range, EPA has recognized the large body of data 

and risk analyses demonstrating that retention of the current standard would 

leave large numbers of individuals at risk for respiratory effects and/or other 

significant health impacts including asthma exacerbations, emergency room 

visits, hospital admissions and mortality. 

In response to EPA’s request for additional advice on the reconsideration in 2011, 

CASAC reaffirmed their conclusion that “the evidence from controlled human and 

epidemiological studies strongly supports the selection of a new primary ozone standard within 

the 60 – 70 ppb range for an 8-hour averaging time” (Samet, 2011). As requested by EPA, 

CASAC’s advice and recommendations were based on the scientific and technical record from 

the 2008 rulemaking. In considering the record for the 2008 rulemaking, CASAC stated the 

following to summarize the basis for their conclusions (Samet, 2011, pp. ii to iii). 

 The evidence available on dose-response for effects of ozone shows 

associations extending to levels within the range of concentrations 

currently experienced in the United States. 
3-126 

 There is scientific certainty that 6.6-hour exposures with exercise of 

young, healthy, non-smoking adult volunteers to concentrations ≥ 80 ppb 

cause clinically relevant decrements of lung function. 

 Some healthy individuals have been shown to have clinically relevant 

responses, even at 60 ppb. 

 Since the majority of clinical studies involve young, healthy adult 

populations, less is known about health effects in such potentially ozone 

sensitive populations as the elderly, children and those with 

cardiopulmonary disease. For these susceptible groups, decrements in 

lung function may be greater than in healthy volunteers and are likely to 

have a greater clinical significance. 

 Children and adults with asthma are at increased risk of acute 

exacerbations on or shortly after days when elevated ozone concentrations 

occur, even when exposures do not exceed the NAAQS concentration of 75 

ppb. 

 Large segments of the population fall into what EPA terms a “sensitive 

population group,’’ i.e., those at increased risk because they are more 

intrinsically susceptible (children, the elderly, and individuals with 

chronic lung disease) and those who are more vulnerable due to increased 

exposure because they work outside or live in areas that are more polluted 

than the mean levels in their communities. 

With respect to evidence from epidemiologic studies, CASAC stated “while epidemiological 

studies are inherently more uncertain as exposures and risk estimates decrease (due to the greater 

potential for biases to dominate small effect estimates), specific evidence in the literature does 

not suggest that our confidence on the specific attribution of the estimated effects of ozone on 

health outcomes differs over the proposed range of 60-70 ppb.” (Samet, 2011, p. 10). 

Following their review of the second draft PA in the current review, which considers an 

updated scientific and technical record since the 2008 rulemaking, CASAC concluded that “there 

is clear scientific support for the need to revise the standard” (Frey, 2014, p. ii). In particular, 

CASAC noted the following (Frey, 2014, p. 5): 

[T]he scientific evidence provides strong support for the occurrence of a range of 

adverse respiratory effects and mortality under air quality conditions that would 

meet the current standard. Therefore, CASAC unanimously recommends that the 

Administrator revise the current primary ozone standard to protect public health. 

In supporting these conclusions, CASAC judged that the strongest evidence comes from 

controlled human exposure studies of respiratory effects. The Committee specifically noted that 

“the combination of decrements in FEV1 together with the statistically significant alterations in 

symptoms in human subjects exposed to 72 ppb ozone meets the American Thoracic Society’s 

definition of an adverse health effect” (Frey, 2014, p. 5). CASAC further judged that “the level at 
3-127 

which adverse effects might be observed would likely be lower for more sensitive subgroups, 

such as those with asthma” (Frey, 2014, p. 5). 

With regard to lung function risk estimates based on information from controlled human 

exposure studies, CASAC concluded that “estimation of FEV1 decrements of ≥15% is 

appropriate as a scientifically relevant surrogate for adverse health outcomes in active healthy 

adults, whereas an FEV1 decrement of ≥10% is a scientifically relevant surrogate for adverse 

health outcomes for people with asthma and lung disease” (Frey, 2014, p. 3). The Committee 

further concluded that “[a]sthmatic subjects appear to be at least as sensitive, if not more 

sensitive, than non-asthmatic subjects in manifesting O3-induced pulmonary function 

decrements” (Frey, 2014, p. 4). In considering estimates of the occurrence of these decrements in 

urban case study areas, CASAC specifically noted that the current standard is estimated to allow 

11 to 22% of school age children to experience at least one day with an FEV1 decrement > 10%. 

While CASAC judged that controlled human exposure studies of respiratory effects 

provide the strongest evidence supporting their conclusion on the current standard, the 

Committee judged that there is also “sufficient scientific evidence based on epidemiologic 

studies for mortality and morbidity associated with short-term exposure to ozone at the level of 

the current standard” (Frey, 2014, p. 5). In support of the biological plausibility of the 

associations reported in these epidemiologic studies, CASAC noted that “[r]ecent animal 

toxicological studies support identification of modes of action and, therefore, the biological 

plausibility associated with the epidemiological findings” (Frey, 2014, p. 5). 

Consistent with the advice of CASAC, several public commenters supported revising the 

primary O3 standard to provide increased public health protection. In considering the available 

evidence as a basis for their views, these commenters generally noted that the health evidence is 

stronger in the current review than in past reviews. These commenters often noted that causal 

determinations were strengthened to “likely causal” for total mortality and cardiovascular effects 

from short-term O3 exposures, and for respiratory effects from long-term O3 exposures. These 

commenters also noted the increase in controlled human exposure studies showing lung function 

decrements and new evidence of inflammation in healthy young adults at 60 ppb O3, as well as 

the increase in the number of epidemiologic studies showing consistent, positive associations 

between O3 exposures and hospital admissions, emergency department visits, and premature 

mortality. Some commenters noted that children have long been known to be more vulnerable 

than adults to the effects of air pollution due to ongoing lung development, the greater 

permeability of their airways epithelial layer, and greater resting minute ventilation (when 

normalized to body mass or lung volume) resulting in increased exposure compared with adults. 

These commenters noted that adverse effects have been described on early lung development and 

the evidence for O3 as a contributor to childhood respiratory disease is extremely strong. They 
3-128 

expressed the view that O3 in particular has long been known to induce asthma exacerbations in 

children and, in one well characterized population-based cohort study in California, exposure to 

ozone was associated with the development of asthma. Some commenters expressed the view 

that young children and small infants should be included in the exposure and risk assessment. 

Other commenters noted that the health endpoints considered in the HREA are limited, and do 

not represent the comprehensive array of health effects attributable to O3 exposure. 

In contrast to the views discussed above, other public commenters opposed considering 

revised standards. These commenters discussed a variety of reasons for their views. A number of 

commenters expressed the view that EPA should not lower the level of the standard because a 

lower level would be closer to background O3 concentrations. In addition, several commenters 

challenged the interpretation of the evidence presented in the ISA. For example, some 

commenters questioned the ISA’s judgments regarding the strength of evidence for 

cardiovascular system effects from short-term O3 exposures. With respect to the risk assessment, 

several commenters expressed the view that the EPA should only estimate risks above O3

background concentrations, or above threshold concentrations. In some cases these commenters 

noted that (1) the O3 mode of action indicates that there are thresholds for O3 effects; (2) that 

these thresholds are considered in the lung function risk assessment; and (3) that there is no 

reason to believe that similar thresholds would not also be associated with other health effects, 

particularly more serious effects. Some commenters also expressed the view that, based on the 

mortality and morbidity risk estimates in the HREA, there is little to no difference between the 

risks estimated for the current O3 standard and the risks estimated for revised standards with 

lower levels. These commenters concluded that the HREA and PA have not shown that the 

public health improvements likely to be achieved by a revised O3 standard would be greater than 

the improvements likely to be achieved by the current standard. 

3.4 STAFF CONCLUSIONS ON ADEQUACY OF PRIMARY STANDARD 

This section presents staff’s conclusions for the Administrator to consider in deciding 

whether it is appropriate to revise the existing primary O3 standard. Staff conclusions are based 

on our consideration of the assessment and integrative synthesis of the evidence presented in the 

ISA, the air quality distributions in locations of selected epidemiologic studies, exposure and risk 

analyses in the HREA, the advice of CASAC, and comments received from members of the 

public. 

As an initial matter, staff concludes that reducing precursor emissions to achieve O3 

concentrations that meet the current standard will provide important improvements in public 

health protection. This initial conclusion is based on (1) the strong body of scientific evidence 

indicating a wide range of adverse health outcomes attributable to exposures to O3
3-129 

concentrations commonly found in the ambient air and (2) estimates indicating decreased 

occurrences of O3 exposures of concern and decreased health risks upon meeting the current 

standard, compared to recent air quality. 

Strong support for this initial conclusion is provided by controlled human exposure 

studies of respiratory effects, and by quantitative estimates of exposures of concern and lung 

function decrements based on information in these studies. Analyses in the HREA estimate that 

the percentages of children (i.e., all children and children with asthma) in urban case study areas 

experiencing exposures of concern, or experiencing abnormal and potentially adverse lung 

function decrements, are consistently lower for air quality that just meets the current O3 standard 

than for recent air quality. The HREA estimates such reductions consistently across the urban 

case study areas evaluated and throughout various portions of individual urban case study areas, 

including in urban cores and the portions of case study areas surrounding urban cores. These 

reductions in exposures of concern and O3-induced lung function decrements reflect the 

consistent decreases in the highest O3 concentrations following reductions in precursor emissions 

to meet the current standard. Thus, populations in both urban and non-urban areas would be 

expected to experience important reductions in O3 exposures and O3-induced lung function risks 

upon meeting the current standard. 

Support for this initial conclusion is also provided by estimates of O3-associated mortality 

and morbidity based on application of concentration-response relationships from epidemiologic 

studies to air quality adjusted to just meet the current standard. These estimates, which are based 

on the assumption that concentration-response relationships are linear over entire distributions of 

ambient O3 concentrations, are associated with uncertainties that complicate their interpretation 

(discussed below). However, risk estimates for effects associated with short- and long-term O3

exposures, combined with the HREA’s national analysis of O3 responsiveness to reductions in 

precursor emissions and the consistent reductions estimated for the highest ambient O3

concentrations, suggest that O3-associated mortality and morbidity would be expected to 

decrease nationwide following reductions in precursor emissions to meet the current O3 standard. 

As discussed in section 3.2.3.2, reductions in O3 precursor emissions (i.e., NOX) could 

also increase public health protection by reducing the ambient concentrations of pollutants other 

than O3. For example, NOX emissions contribute to ambient NO2, and NOX and VOCs can 

contribute to secondary formation of PM2.5 constituents, including ammonium sulfate (NH4SO4), 

ammonium nitrate (NH4NO3), and organic carbon (OC). Therefore, at some times and in some 

locations, control strategies that would reduce NOX emissions (i.e., to meet an O3 standard) could 

reduce ambient concentrations of NO2 and PM2.5, resulting in health benefits beyond those 

directly associated with reducing ambient O3 concentrations. 
3-130 

We next revisit the overarching policy question for this chapter, taking into consideration 

the responses to the specific questions focused on the adequacy of the current primary O3

standard, as discussed above. 

 Does the currently available scientific evidence and exposure/risk information, 

as reflected in the ISA and HREA, support or call into question the adequacy 

of the protection afforded by the current primary O3 standard? 

In considering the available evidence and information, staff concludes that the O3-

attributable health effects estimated to be allowed by air quality that meets the current primary 

standard for O3 can reasonably be judged important from a public health perspective. Thus, we 

conclude that the available health evidence and exposure/risk information call into question the 

adequacy of the public health protection provided by the current standard. We further conclude 

that it is appropriate in this review to consider alternative standards that would increase public 

health protection, compared to the current standard. The basis for these conclusions is discussed 

below. 

Studies evaluated since the completion of the 2006 O3 AQCD support and expand upon 

the strong body of evidence that, in the last review, indicated a causal relationship between shortterm O3 exposures and respiratory health effects. Together, experimental and epidemiologic 

studies support conclusions regarding a continuum of O3 respiratory effects ranging from small 

reversible changes in pulmonary function to more serious effects that can result in respiratoryrelated emergency department visits, hospital admissions, and/or mortality. Recent animal 

toxicological studies support descriptions of modes of action for these respiratory effects and 

augment support for biological plausibility for the role of O3 in reported effects. With regard to 

mode of action, evidence indicates that antioxidant capacity may modify the risk of respiratory 

morbidity associated with O3 exposure. In addition, based on the consistency of findings across 

studies and evidence for the coherence of results from different scientific disciplines, strong 

evidence indicates that certain populations are at increased risk of experiencing O3-related 

effects. These include populations and lifestages identified in previous reviews (i.e., people with 

asthma, children, older adults, outdoor workers) and populations identified since the last review 

(i.e., people with certain genotypes related to anti-oxidant and/or anti-inflammatory status; 

people with reduced intake of certain nutrients, such as Vitamins C and E). 

Evidence for adverse respiratory health effects attributable to “long-term” or repeated 

daily O3 exposures is much stronger than in previous reviews, and the ISA concludes that the 

evidence supports a “likely to be” causal relationship between such O3 exposures and adverse 

respiratory health effects. Uncertainties related to the extrapolation of data generated by rodent 

toxicology studies to the understanding of health effects in humans have been reduced by studies 
3-131 

in non-human primates and by recent epidemiologic studies. The evidence available in this 

review includes new epidemiologic studies using a variety of designs and analysis methods, 

conducted by different research groups in different locations, evaluating the relationships 

between long-term O3 exposures and measures of respiratory morbidity and mortality. New 

evidence supports associations between long-term or repeated O3 exposures and the development 

of asthma in children, with several studies reporting interactions between genetic variants and 

such O3 exposures. Studies also report associations between long-term or repeated O3 exposure 

and asthma prevalence, asthma severity and control, respiratory symptoms among asthmatics, 

and respiratory mortality. 

In considering the O3 exposure concentrations reported to elicit respiratory effects, we 

note that controlled human exposure studies provide the most certain evidence indicating the 

occurrence of health effects in humans following exposures to specific O3 concentrations. 

Consistent with this, CASAC also concluded that “the scientific evidence supporting the finding 

that the current standard is inadequate to protect public health is strongest based on the 

controlled human exposure studies of respiratory effects” (Frey, 2014, p. 5). As discussed above, 

recent evidence includes controlled human exposure studies reporting lung function decrements 

and pulmonary inflammation in healthy adults engaged in intermittent, moderate exertion 

following 6.6 hour exposures to O3 concentrations as low as 60 ppb, and lung function 

decrements and respiratory symptoms following exposures to concentrations as low as 72 ppb.97

Compared to the evidence available in the last review, these studies have strengthened support 

for the occurrence of abnormal and adverse respiratory effects attributable to short-term 

exposures to O3 concentrations below 80 ppb.98 Consistent with CASAC advice, we conclude 

that exposures to such O3 concentrations are potentially important from a public health 

perspective given the following: 

1. The respiratory effects reported to occur in healthy adults following exposures to O3

concentrations of 60 and 72 ppb, while at moderate exertion, can reasonably be judged 

adverse based on ATS criteria and advice from CASAC. In considering the 72 ppb 

exposure concentration, CASAC noted that “the combination of decrements in FEV1

together with the statistically significant alterations in symptoms in human subjects 

exposed to 72 ppb ozone meets the American Thoracic Society’s definition of an adverse 

health effect” (Frey, 2014, p. 5). With regard to 60 ppb O3, CASAC agreed that “a level 

of 60 ppb corresponds to the lowest exposure concentration demonstrated to result in 

 

97 As noted above, for the 70 ppb exposure concentration Schelegle et al. (2009) reported that the actual mean 

exposure concentration was 72 ppb. 

98 Cf. Misisssippi. 744 F.3d at 1350 (“Perhaps more studies like the Adams studies will yet reveal that the 0.060 ppm 

level produces significant adverse decrements that simply cannot be attributed to normal variation in lung 

function.”). 
3-132 

lung function decrements large enough to be judged an abnormal response by ATS and 

that could be adverse in individuals with lung disease” (Frey, 2014, p. 7). CASAC further 

noted that “a level of 60 ppb also corresponds to the lowest exposure concentration at 

which pulmonary inflammation has been reported” (Frey, 2014, p. 7). 

2. The controlled human exposure studies reporting these respiratory effects were conducted 

in healthy adults, while at-risk groups (e.g., children, people with asthma) could 

experience larger and/or more serious effects. In their advice to the Administrator, 

CASAC concurred with this conclusion (Frey, 2014, p. 5). 

3. These respiratory effects are coherent with the serious health outcomes that have been 

reported in epidemiologic studies (e.g., respiratory-related hospital admissions, 

emergency department visits, and mortality). 

Given the above considerations, our conclusions regarding the adequacy of the current primary 

O3 standard place a large amount of weight on the results of controlled human exposure studies 

conducted at 60 and 72 ppb, and on HREA analyses based on information from controlled 

human exposure studies (i.e., exposures of concern to O3 concentrations at or above 60, 70, and 

80 ppb and O3-induced FEV1 decrements ≥ 10%, 15%, and 20%). 

Recent epidemiologic studies also provide support, beyond that available in the last 

review, for associations between short-term O3 exposures and a wide range of adverse 

respiratory outcomes (including respiratory-related hospital admissions, emergency department 

visits, and mortality) and with total mortality. Associations with morbidity and mortality are 

stronger during the warm or summer months, and remain robust after adjustment for copollutants. Many epidemiologic studies of morbidity effects and mortality were conducted in 

locations that did not meet the current standard. However, in one U.S. single-city study 

associations with respiratory morbidity were reported in a location that would likely have met the 

current O3 standard over the entire study period, suggesting that health effect associations persist 

in locations meeting the current standard. In addition, associations with respiratory morbidity or 

mortality were reported in several Canadian multicity studies, and in cut point analyses included 

in a U.S. multicity study, when the majority of study locations would likely have met the current 

O3 standard. While there is additional uncertainty in interpreting the relationship between air 

quality meeting the current standard and health effects in these multicity studies (i.e., compared 

to single-city studies), they provide supporting evidence for the occurrence of health effect 

associations in locations that meet the current standard. Even in some study locations where the 

current standard was likely not met, considering reported concentration-response functions in the 

context of available air quality data support the occurrence of O3-health effect associations on the 

subsets of days with ambient O3 concentrations below the level of the current standard. Taken 

together, these studies and associated air quality data support the occurrence of O3-associated 
3-133 

hospital admissions, emergency department visits, and mortality at ambient concentrations that 

meet the current standard. 

Beyond our consideration of the evidence, we also consider the results of the HREA 

exposure and risk analyses in reaching conclusions regarding the adequacy of the current 

primary O3 standard. In doing so, we focus primarily on estimates of the occurrence of exposures 

of concern to O3 concentrations at or above 60 and 70 ppb and lung function decrements > 10%, 

15% and 20%. We place relatively less weight on epidemiologic-based risk estimates, noting that 

the overall conclusions from the HREA likewise reflect less confidence in estimates of 

epidemiologic-based risks than in estimates of exposures and lung function risks (U.S. EPA, 

2014, section 9.6). Our determination to attach less weight to the epidemiologic-based estimates 

reflects the uncertainties associated with mortality and morbidity risk estimates, including the 

heterogeneity in effect estimates between locations, the potential for exposure measurement 

errors, and uncertainty in the interpretation of the shape of concentration-response functions at 

lower O3 concentrations. The HREA also concludes that lower confidence should be placed in 

the results of the assessment of respiratory mortality risks associated with long-term O3

exposures, primarily because that analysis is based on only one study (even though that study is 

well-designed) and because of the uncertainty in that study about the existence and level of a 

potential threshold in the concentration-response function (U.S. EPA, 2014, section 9.6). 

With regard to HREA estimates of exposures of concern we note the CASAC conclusion 

that 60 ppb is an appropriate exposure of concern for asthmatic children (Frey, 2014, p. 8). 

Exposure estimates from the HREA indicate that, if the 15 urban case study areas were to just 

meet the current O3 standard, approximately 10 to 20% of children (on average over the years of 

analysis) in those areas, including asthmatic children, could experience one or more exposures of 

concern to O3 concentrations of 60 ppb or above. In the case study areas evaluated, this 

corresponds to over 2 million children (including over 200,000 asthmatic children) experiencing 

approximately 4 million such exposures. Nationally, far more children would be expected to 

experience such exposures of concern. On average over the years evaluated in the HREA, 

approximately 3 to 8% of children are estimated to experience two or more exposures of concern 

to O3 concentrations of 60 ppb or greater. For the worst-case years in the worst-case locations 

(i.e., years and locations with air quality patterns resulting in the largest exposure estimates), 

approximately 25% of children are estimated to experience one or more exposures of concern at 

or above 60 ppb, and about 14% are estimated to experience two or more such exposures. 

Although the current standard more effectively limits exposures of concern at or above higher O3

concentrations (i.e., 70, 80 ppb), we note that in the worst-case year and location about 8% of 

children are estimated to experience one or more exposures of concern at or above 70 ppb and 

about 2% of children are estimated to experience two or more such exposures. 
3-134 

Though we focus on children in these analyses of O3 exposures, we also recognize that 

exposures to 8-hour average O3 concentrations at or above 60, 70, or 80 ppb could be of concern 

for adult populations as well. As discussed above, the patterns of exposure estimates over years 

and across cities are similar in adult asthmatics, older adults, and children, though smaller 

percentages of adult populations are estimated to experience exposures of concern. Thus, the 

results for children are one part of a broader range of at-risk populations that also includes 

asthmatic adults and older adults. 

Consistent with estimates of exposures of concern, the HREA also estimates that under 

air quality conditions just meeting the current O3 NAAQS, hundreds of thousands of asthmatic 

children would be expected to experience O3-induced lung function decrements that are large 

enough to be potentially adverse in people with lung disease. On average over the years 

evaluated in the HREA, the current standard is estimated to allow about 14% to 19% of children 

in the 15 urban case study areas, including asthmatic children, to experience one or more O3-

induced lung function decrements > 10% (a decrement judged by CASAC to be a “scientificallyrelevant surrogate for adverse health outcomes for people with asthma and lung disease” (Frey, 

2014, p. 4)). This corresponds to about 300,000 asthmatic children. Nationally, far more children 

would be expected to experience such O3-induced lung function decrements. Across the 15 urban 

areas, about 8% to 12% of children are estimated to experience two or more decrements > 10%, 

on average over the analysis years. In the worst-case year and location, approximately 22% of 

children are estimated to experience one or more decrements > 10% and about 14% are 

estimated to experience two or more such decrements. As with exposures of concern, the current 

standard more effectively limits the larger O3-induced lung function decrements evaluated (i.e., > 

15%, 20%). However, about 7% of children are estimated to experience one or more O3-induced 

decrements > 15% in the worst-case city and year analyzed in the HREA, and about 4% are 

estimated to experience two or more such decrements. As discussed above, CASAC judged 

decrements > 15% to be an appropriate “surrogate for adverse health outcomes in active healthy 

adults” (Frey, 2014, p. 4). 

As noted above, compared to the weight given to HREA estimates of exposures of 

concern and lung function risks, we place relatively less weight on epidemiologic-based risk 

estimates. For epidemiology-based risk estimates, we consider total risks (i.e., based on the full 

distributions of ambient O3 concentrations) and risks associated with O3 concentrations in the 

upper portions of ambient distributions. A focus on estimates of total risks places greater weight 

on the possibility that concentration-response relationships remain linear over the entire 

distributions of ambient O3 concentrations. With regard to total risks, the HREA estimates 

thousands of O3-associated hospital admissions, emergency department visits, and deaths per 

year for air quality conditions associated with just meeting the current standard in the 12 urban 
3-135 

case study areas evaluated. A focus on risks associated with O3 concentrations in the upper 

portions of ambient distributions places greater weight on the uncertainty associated with the 

shapes of concentration-response curves for O3 concentrations in the lower portions of ambient 

distributions (section 3.2.3.2). Based on area-wide O3 concentrations from the upper portions of 

seasonal distributions, the current standard is estimated to allow hundreds to thousands of O3-

associated deaths per year in urban case study areas. As with the exposures of concern and lung 

function risks, this number would be much greater if risks were assessed across the entire U.S. 

population. 

Although we note the HREA conclusions indicating somewhat less confidence in 

estimates of O3-associated mortality and morbidity risks compared to estimates of exposures of 

concern and lung function risks, we conclude that the general magnitude of mortality and 

morbidity risk estimates suggests the potential for a substantial number of O3-associated deaths 

and adverse respiratory events nationally when the current standard is met. This is the case even 

based on the risks associated with the upper ends of distributions of ambient O3 concentrations, 

where experimental evidence indicates increasing support for the occurrence of adverse effects 

attributable to O3 exposures. 

In addition to the evidence and exposure/risk information discussed above, we also take 

note of the CASAC advice provided to the EPA Administrator on the proposed reconsideration 

of the 2008 decision establishing the current standard and the advice of CASAC in the current 

review. In commenting on the proposed reconsideration, the prior CASAC O3 Panel 

recommended revision of the standard to one with a lower level based on the evidence and 

information in the record for the 2008 standard (Samet, 2011), which has been substantially 

strengthened in the current review. As discussed in more detail above, the current CASAC also 

“unanimously recommends that the Administrator revise the current primary ozone standard to 

protect public health” (Frey, 2014, p. 6). 

In consideration of all of the above, staff reaches the conclusion that the available 

evidence and exposure and risk information clearly calls into question the adequacy of public 

health protection provided by the current primary standard. The evidence from controlled human 

exposure studies provides strong support for the occurrence of adverse respiratory effects 

following exposures to O3 concentrations below the level of the current standard. Epidemiologic 

studies provide support for the occurrence of adverse respiratory effects and mortality under air 

quality conditions that would likely meet the current standard. In addition, based on the analyses 

in the HREA, we conclude that the exposures and risks projected to remain upon meeting the 

current standard are indicative of risks that can reasonably be judged to be important from a 

public health perspective. Thus, staff concludes that the evidence and information provides 

strong support for giving consideration to revising the current primary standard in order to 
3-136 

provide increased public health protection against an array of adverse health effects that range 

from decreased lung function and respiratory symptoms to more serious indicators of morbidity 

(e.g., including emergency department visits and hospital admissions), and mortality. In 

consideration of all of the above, staff draws the conclusion that it is appropriate for the 

Administrator to consider revision of the current primary O3 standard to provide increased public 

health protection. 
3-137 

3.5 REFERENCES 

Adams, W. C. (2006) Comparison of chamber 6.6 hour exposures to 0.04-0.08 ppm ozone via square-wave and 

triangular profiles on pulmonary responses. Inhalation Toxicol. 18: 127-136. 

http://dx.doi.org/10.1080/08958370500306107 

Adams, WC. (2003). Comparison of chamber and face mask 6.6-hour exposure to 0.08 ppm ozone via square-wave 

and triangular profiles on pulmonary responses. Inhal Toxicol 15: 265-281. 

Adams, WC. (2002). Comparison of chamber and face-mask 6.6-hour exposures to ozone on pulmonary function 

and symptoms responses. Inhal Toxicol 14: 745-764. http://dx.doi.org/10.1080/08958370290084610 

Adams, WC. (1998). Dose-response effect of varied equivalent minute ventilation rates on pulmonary function 

responses during exposure to ozone. Washington, DC: American Petroleum Institute. 

Adams, P. F.; Hendershot, G. E.; Marano, M. A. (1999) Current estimates from the National Health Interview 

Survey, 1996. Hyattsville, MD: U.S. Department of Health and Human Services, Public Health Service, 

National Center for Health Statistics; DHHS publication no. (PHS) 99-1528. (Vital and health statistics: v. 

10, data from the National Health Survey, no. 200). Available: 

http://www.cdc.gov/nchs/products/pubs/pubd/series/sr10/pre-200/pre-200.htm [12 March, 2001]. 

Adamson, I; Prieditis, H. (1995). Response of mouse lung to carbon deposition during injury and repair. 

Environmental Health Perspectives. 103: 1: 72-76. 

Akinbami, LJ; Lynch, CD; Parker, JD; Woodruff, TJ. (2010). The association between childhood asthma prevalence 

and monitored air pollutants in metropolitan areas, United States, 2001-2004. Environ Res 110: 294-301. 

http://dx.doi.org/10.1016/j.envres.2010.01.001 

Alexeeff, SE; Litonjua, AA; Suh, H; Sparrow, D; Vokonas, PS; Schwartz, J. (2007). Ozone exposure and lung 

function: Effect modified by obesity and airways hyperresponsiveness in the VA Normative Aging Study. 

Chest 132: 1890-1897. http://dx.doi.org/10.1378/chest.07-1126 

Alexis, NE; Lay, JC; Hazucha, M; Harris, B; Hernandez, ML; Bromberg, PA; Kehrl, H; Diaz-Sanchez, D; Kim, C; 

Devlin, RB; Peden, DB. (2010). Low-level ozone exposure induces airways inflammation and modifies cell 

surface phenotypes in healthy humans. Inhal Toxicol 22: 593-600. 

Alexis, N; Urch, B; Tarlo, S; Corey, P; Pengelly, D; O'Byrne, P; Silverman, F. (2000). Cyclooxygenase metabolites 

play a different role in ozone-induced pulmonary function decline in asthmatics compared to normals. Inhal 

Toxicol 12: 1205-1224. 

Aris, RM; Tager, I; Christian, D; Kelly, T; Balmes, JR. (1995). Methacholine responsiveness is not associated with 

O3-induced decreases in FEV1. Chest 107: 621-628. 

ATS (American Thoracic Society). (2000). What constitutes an adverse health effect of air pollution? Am. J. Respir. 

Crit. Care Med. 161: 665-673. 

ATS (American Thoracic Society). (1991). Lung function testing: selection of reference values and interpretive 

strategies. Am. Rev. Respir. Dis. 144: 1202-1218. 

ATS (American Thoracic Society). (1985) Guidelines as to what constitutes an adverse respiratory health effect, 

with special reference to epidemiological studies of air pollution. Am. Rev. Respir. Dis. 131: 666-668. 

Avol, EL; Trim, SC; Little, DE; Spier, CE; Smith, MN; Peng, RC; Linn, WS; Hackney, JD; Gross, KB; D'Arcy, JB; 

Gibbons, D; Higgins, ITT. (1990). Ozone exposure and lung function in children attending a southern 
3-138 

California summer camp. In Proceedings of the 83rd A&WMA Annual Meeting. Pittsburgh, PA: Air & 

Waste Management Association. 

Balmes, JR; Chen, LL; Scannell, C; Tager, I; Christian, D; Hearne, PQ; Kelly, T; Aris, RM. (1996). Ozoneinduced 

decrements in FEV1 and FVC do not correlate with measures of inflammation. Am J Respir Crit Care Med 

153: 904-909. 

Barraza-Villarreal, A; Sunyer, J; Hernandez-Cadena, L; Escamilla-Nunez, MC; Sienra-Monge, JJ; Ramirez-Aguilar, 

M; Cortez-Lugo, M; Holguin, F; Diaz-Sanchez, D; Olin, AC; Romieu, I. (2008). Air pollution, airway 

inflammation, and lung function in a cohort study of Mexico City schoolchildren. Environ Health Perspect 

116: 832-838. http://dx.doi.org/10.1289/ehp.10926 

Basha, MA; Gross, KB; Gwizdala, CJ; Haidar, AH; Popovich, J, Jr. (1994). Bronchoalveolar lavage neutrophilia in 

asthmatic and healthy volunteers after controlled exposure to ozone and filtered purified air. Chest 106: 

1757-1765. 

Bell, ML; Peng, RD; Dominici, F. (2006). The exposure-response curve for ozone and risk of mortality and the 

adequacy of current ozone regulations. Environ Health Perspect 114: 532-536. 

Berhane, K; Zhang, Y; Linn, WS; Rappaport, EB; Bastain, TM; Salam, MT; Islam, T; Lurmann, F; Gilliland, FD. 

(2011). The effect of ambient air pollution on exhaled nitric oxide in the Children's Health Study. Eur 

Respir J 37: 1029-1036. http://dx.doi.org/10.1183/09031936.00081410 

Berry, M; Lioy, PJ; Gelperin, K; Buckler, G; Klotz, J. (1991). Accumulated exposure to ozone and measurement of 

health effects in children and counselors at two summer camps. Environ Res 54: 135-150. 

Blomberg, A; Mudway, IS; Nordenhall, C; Hedenstrom, H; Kelly, FJ; Frew, AJ; Holgate, ST; Sandstrom, T. (1999). 

Ozone-induced lung function decrements do not correlate with early airway inflammatory or antioxidant 

responses. Eur Respir J 13: 1418-1428. 

Bloom B, Cohen RA, Freeman G. (2011). Summary health statistics for U.S. children: National Health Interview 

Survey, 2010. National Center for Health Statistics. Vital Health Stat 10(250). 

http://www.cdc.gov/nchs/data/series/sr_10/sr10_250.pdf 

Bosson, J; Stenfors, N; Bucht, A; Helleday, R; Pourazar, J; Holgate, ST; Kelly, FJ; Sandstrom, T; Wilson, S; Frew, 

AJ; Blomberg, A. (2003). Ozone-induced bronchial epithelial cytokine expression differs between healthy 

and asthmatic subjects. Clin Exp Allergy 33: 777-782. 

Brauer, M; Blair, J; Vedal, S. (1996). Effect of ambient ozone exposure on lung function in farm workers. Am J 

Respir Crit Care Med 154: 981-987. 

Breton, CV; Salam, MT; Vora, H; Gauderman, WJ; Gilliland, FD. (2011). Genetic variation in the glutathione 

synthesis pathway, air pollution, and children's lung function growth. Am J Respir Crit Care Med 183: 243-

248. http://dx.doi.org/10.1164/rccm.201006-0849OC 

Brown, JS; Bateson, TF; McDonnell, WF. (2008). Effects of exposure to 0.06 ppm ozone on FEV1 in humans: A 

secondary analysis of existing data. Environ Health Perspect 116: 1023-1026. 

http://dx.doi.org/10.1289/ehp.11396 

Brunekreef, B; Hoek, G; Breugelmans, O; Leentvaar, M. (1994). Respiratory effects of low-level photochemical air 

pollution in amateur cyclists. Am J Respir Crit Care Med 150: 962-966. 

Cakmak, S; Dales, RE; Angelica Rubio, M; Blanco Vidal, C. (2011). The risk of dying on days of higher air 

pollution among the socially disadvantaged elderly. Environ Res 111: 388-393. 

http://dx.doi.org/10.1016/j.envres.2011.01.003 
3-139 

Cakmak, S; Dales, RE; Vidal, CB. (2007). Air pollution and mortality in Chile: Susceptibility among the elderly. 

Environ Health Perspect 115: 524-527. 

Cakmak, S; Dales, RE; Judek, S. (2006). Respiratory health effects of air pollution gases: Modification by education 

and income. Arch Environ Occup Health 61: 5-10. http://dx.doi.org/10.3200/AEOH.61.1.5-10 

Carey, SA; Minard, KR; Trease, LL; Wagner, JG; Garcia, GJ; Ballinger, CA; Kimbell, JS; Plopper, CG; Corley, 

RA; Postlethwait, EM; Harkema, JR; Einstein, DR. (2007). Three-dimensional mapping of ozone-induced 

injury in the nasal airways of monkeys using magnetic resonance imaging and morphometric techniques. 

Toxicol Pathol 35: 27-40. http://dx.doi.org/10.1080/01926230601072343 

Castillejos, M; Gold, DR; Damokosh, AI; Serrano, P; Allen, G; McDonnell, WF; Dockery, D; Velasco, SR; 

Hernandez, M; Hayes, C. (1995). Acute effects of ozone on the pulmonary function of exercising 

schoolchildren from Mexico City. Am J Respir Crit Care Med 152: 1501-1507. 

Chan, CC; Wu, TH. (2005). Effects of ambient ozone exposure on mail carriers' peak expiratory flow rates. Environ 

Health Perspect 113: 735-738. http://dx.doi.org/10.1289/ehp.7636 

Chen, J; Tan, M; Nemmar, A; Song, W; Dong, M; Zhang, G; Li, Y. (2006). Quantification of extrapulmonary 

translocation of intratracheal-instilled particles in vivo in rats: effect of lipopolysaccharide. Toxicology 

222:195-201. 

Chhabra, SK; Yasir, A; Chaudhry, K; Shah, B. (2010). Effect of ozone on response to ovalbumin & its modulation 

by vitamins C & E in sensitized guinea pigs. Indian J Med Res 132: 87-93. 

Cockcroft, DW; Davis, BE; Todd, DC; Smycniuk, AJ. (2005). Methacholine challenge: Comparison of two 

methods. Chest 127: 839-844. 

Cohen, MD; Sisco, M; Baker, K; Li, Y; Lawrence, D; Van Loveren, H; Zelikoff, JT; Schlesinger, RB. (2002). 

Effects of inhaled ozone on pulmonary immune cells critical to antibacterial responses in situ. Inhal 

Toxicol 14: 599-619. http://dx.doi.org/10.1080/08958370290084520 

Dales, RE; Cakmak, S; Doiron, MS. (2006). Gaseous air pollutants and hospitalization for respiratory disease in the 

neonatal period. Environ Health Perspect 114: 1751-1754. http://dx.doi.org/10.1289/ehp.9044 

Delfino, RJ; Staimer, N; Tjoa, T; Arhami, M; Polidori, A; Gillen, DL; George, SC; Shafer, MM; Schauer, JJ; 

Sioutas, C. (2010a). Associations of primary and secondary organic aerosols with airway and systemic 

inflammation in an elderly panel cohort. Epidemiology 21: 892-902. 

http://dx.doi.org/10.1097/EDE.0b013e3181f20e6c 

DeNavas-Walt, Carmen, Bernadette D. Proctor, and Jessica C. Smith. (2011) U.S. Census Bureau, Current 

Population Reports, P60-239, Income, Poverty, and Health Insurance Coverage in the United States: 2010, 

U.S. Government Printing Office, Washington, DC. http://www.census.gov/prod/2011pubs/p60-239.pdf 

Depuydt, P; Joos, GF; Pauwels, RA. (1999). Ambient ozone concentrations induce airway hyperresponsiveness in 

some rat strains. Eur Respir J 14: 125-131. 

Devlin, RB; Mcdonnell, WF; Mann, R; Becker, S; House, DE; Schreinemachers, D; Koren, HS. (1991). Exposure of 

humans to ambient levels of ozone for 6.6 hours causes cellular and biochemical changes in the lung. Am J 

Respir Cell Mol Biol 4: 72-81. 

Devlin, RB; Folinsbee, LJ; Biscardi, F; Hatch, G; Becker, S; Madden, MC; Robbins, M; Koren, HS. (1997). 

Inflammation and cell damage induced by repeated exposure of humans to ozone. Inhal Toxicol 9: 211-

235. 
3-140 

Dimeo, MJ; Glenn, MG; Holtzman, MJ; Sheller, JR; Nadel, JA; Boushey, HA. (1981). Threshold concentration of 

ozone causing an increase in bronchial reactivity in humans and adaptation with repeated exposures. Am 

Rev Respir Dis 124: 245-248. 

Driscoll, KE; Vollmuth, TA; Schlesinger, RB. (1987). Acute and subchronic ozone inhalation in the rabbit: 

Response of alveolar macrophages. J Toxicol Environ Health 21: 27-43. 

http://dx.doi.org/10.1080/15287398709531000 

Dryden, DM; Spooner, CH; Stickland, MK; Vandermeer, B; Tjosvold, L; Bialy, L; Wong, K; Rowe, BH. (2010). 

Exercise-induced bronchoconstriction and asthma. (AHRQ Publication No. 10-E001). Rockville, MD: 

Agency for Healthcare Research and Quality. 

Escamilla-Nuñez, MC; Barraza-Villarreal, A; Hernandez-Cadena, L; Moreno-Macias, H; Ramirez-Aguilar, M; 

Sienra-Monge, JJ; Cortez-Lugo, M; Texcalac, JL; del Rio-Navarro, B; Romieu, I. (2008). Traffic-related 

air pollution and respiratory symptoms among asthmatic children, resident in Mexico City: The EVA 

cohort study. Respir Res 9: 74. http://dx.doi.org/10.1186/1465-9921-9-74 

Fanucchi, MV; Plopper, CG; Evans, MJ; Hyde, DM; Van Winkle, LS; Gershwin, LJ; Schelegle, ES. (2006). Cyclic 

exposure to ozone alters distal airway development in infant rhesus monkeys. Am J Physiol Lung Cell Mol 

Physiol 291: L644-L650. http://dx.doi.org/10.1152/ajplung.00027.2006 

Feo Brito, F; Mur Gimeno, P; Martinez, C; Tobias, A; Suarez, L; Guerra, F; Borja, JM; Alonso, AM. (2007). Air 

pollution and seasonal asthma during the pollen season: A cohort study in Puertollano and Ciudad Real 

(Spain). Allergy 62: 1152-1157. http://dx.doi.org/10.1111/j.1398-9995.2007.01438.x 

Folinsbee, LJ; McDonnell, WF; Horstman, DH. (1988). Pulmonary function and symptom responses after 6.6-hour 

exposure to 0.12 ppm ozone with moderate exercise. J Air Waste Manag Assoc 38: 28-35. 

Folinsbee, LJ; Hazucha, MJ. (2000). Time course of response to ozone exposure in healthy adult females. Inhal 

Toxicol 12: 151-167. 

Folinsbee, LJ; Hazucha, MJ. (1989). Persistence of ozone-induced changes in lung function and airway 

responsiveness. In Atmospheric ozone research and its policy implications. Amsterdam, The Netherlands: 

Elsevier. 

Folinsbee, LJ; Horstman, DH; Kehrl, HR; Harder, S; Abdul-Salaam, S; Ives, PJ. (1994). Respiratory responses to 

repeated prolonged exposure to 0.12 ppm ozone. Am J Respir Crit Care Med 149: 98-105. 

Frampton, MW; Morrow, PE; Torres, A; Cox, C; Voter, KZ; Utell, MJ; Gibb, FR; Speers, DM. (1997). Ozone 

responsiveness in smokers and nonsmokers. Am J Respir Crit Care Med 155: 116-121. 

Franze, T; Weller, MG; Niessner, R; Pöschl, U. (2005). Protein nitration by polluted air. Environ Sci Technol 39: 

1673-1678. http://dx.doi.org/10.1021/es0488737 

Frey, HC. CASAC Review of the EPA’s Second Draft Policy Assessment for the Review of the Ozone National 

Ambient Air Quality Standards EPA-CASAC-14-004. June 26, 2014. Available online at: 

http://yosemite.epa.gov/sab/sabproduct.nsf/264cb1227d55e02c85257402007446a4/5EFA320CCAD326E8

85257D030071531C/$File/EPA-CASAC-14-004+unsigned.pdf 

Gauderman, WJ; Avol, E; Gilliland, F; Vora, H; Thomas, D; Berhane, K; McConnell, R; Kuenzli, N; Lurmann, F; 

Rappaport, E; Margolis, H; Bates, D; Peters, J. (2004). The effect of air pollution on lung development 

from 10 to 18 years of age. N Engl J Med 351: 1057-1067. 

Gershwin, LJ; Osebold, JW; Zee, YC. (1981). Immunoglobulin E-containing cells in mouse lung following allergen 

inhalation and ozone exposure. International Arch Allergy Appl Immunol 65: 266-277. 
3-141 

Gielen, MH; Van Der Zee, SC; Van Wijnen, JH; Van Steen, CJ; Brunekreef, B. (1997). Acute effects of summer air 

pollution on respiratory health of asthmatic children. Am J Respir Crit Care Med 155: 2105-2108. 

Gold, DR; Damokosh, AI; III, PC; Dockery, DW; McDonnell, WF; Serrano, P; Retama, A; Castillejos, M. (1999). 

Particulate and ozone pollutant effects on the respiratory function of children in southwest Mexico City. 

Epidemiology 10: 8-16. 

Gong, H, Jr; McManus, MS; Linn, WS. (1997). Attenuated response to repeated daily ozone exposures in asthmatic 

subjects. Arch Environ Occup Health 52: 34-41. 

Greer, JR; Abbey, DE; Burchette, RJ. (1993). Asthma related to occupational and ambient air pollutants in 

nonsmokers. J Occup Environ Med 35: 909-915. 

Harkema, JR; Plopper, CG; Hyde, DM; St George, JA; Dungworth, DL. (1987). Effects of an ambient level of ozone 

on primate nasal epithelial mucosubstances: quantitative histochemistry. Am J Pathol 127: 90-96. 

Hatch, GE; Slade, R; Harris, LP; Mcdonnell, WF; Devlin, RB; Koren, HS; Costa, DL; Mckee, J. (1994). Ozone dose 

and effect in humans and rats: A comparison using oxygen-18 labeling and bronchoalveolar lavage. Am J 

Respir Crit Care Med 150: 676-683. 

Henderson, R. (2008) Letter from CASAC Chairman Rogene Henderson to EPA Administrator Stephen Johnson. 

April 7, 2008, EPA-CASAC-08-001. 

Henderson, R. (2006) Letter from CASAC Chairman Rogene Henderson to EPA Administrator Stephen Johnson. 

October 24, 2006, EPA-CASAC-07-001. 

Hernandez, ML; Lay, JC; Harris, B; Esther, CR; Brickey, WJ; Bromberg, PA; Diaz-Sanchez, D; Devlin, RB; 

Kleeberger, SR; Alexis, NE; Peden, DB. (2010). Atopic asthmatic subjects but not atopic subjects without 

asthma have enhanced inflammatory response to ozone. J Allergy Clin Immunol 126: 537-544. 

http://dx.doi.org/10.1016/j.jaci.2010.06.043 

Hernández-Cadena, L; Holguin, F; Barraza-Villarreal, A; Del Río-Navarro, BE; Sienra-Monge, JJ; Romieu, I. 

(2009). Increased levels of outdoor air pollutants are associated with reduced bronchodilation in children 

with asthma. Chest 136: 1529-1536. http://dx.doi.org/10.1378/chest.08-1463 

Halonen, JI; Lanki, T; Tiittanen, P; Niemi, JV; Loh, M; Pekkanen, J. (2009). Ozone and cause-specific 

cardiorespiratory morbidity and mortality. J Epidemiol Community Health 64: 814-820. 

http://dx.doi.org/10.1136/jech.2009.087106 

Hiltermann, JTN; Lapperre, TS; Van Bree, L; Steerenberg, PA; Brahim, JJ; Sont, JK; Sterk, PJ; Hiemstra, PS; Stolk, 

J. (1999). Ozone-induced inflammation assessed in sputum and bronchial lavage fluid from asthmatics: A 

new noninvasive tool in epidemiologic studies on air pollution and asthma. Free Radic Biol Med 27: 1448-

1454. 

Hoek, G; Brunekreef, B; Kosterink, P; Van den Berg, R; Hofschreuder, P. (1993). Effect of ambient ozone on peak 

expiratory flow of exercising children in the Netherlands. Arch Environ Occup Health 48: 27-32. 

http://dx.doi.org/10.1080/00039896.1993.9938390 

Holz, O; Jorres, RA; Timm, P; Mucke, M; Richter, K; Koschyk, S; Magnussen, H. (1999). Ozone-induced airway 

inflammatory changes differ between individuals and are reproducible. Am J Respir Crit Care Med 159: 

776-784. 

Holz, O; Mucke, M; Paasch, K; Bohme, S; Timm, P; Richter, K; Magnussen, H; Jorres, RA. (2002). Repeated ozone 

exposures enhance bronchial allergen responses in subjects with rhinitis or asthma. Clin Exp Allergy 32: 

681-689. 
3-142 

Holz, O; Tal-Singer, R; Kanniess, F; Simpson, KJ; Gibson, A; Vessey, RSJ; Janicki, S; Magnussen, H; Jorres, RA; 

Richter, K. (2005). Validation of the human ozone challenge model as a tool for assessing 

antiinflammatory drugs in early development. J Clin Pharmacol 45: 498-503. 

Hoppe, P; Peters, A; Rabe, G; Praml, G; Lindner, J; Jakobi, G; Fruhmann, G; Nowak, D. (2003). Environmental 

ozone effects in different population subgroups. Int J Hyg Environ Health 206: 505-516. 

http://dx.doi.org/10.1078/1438-4639-00250 

Horstman, DH; Ball, BA; Brown, J; Gerrity, T; Folinsbee, LJ. (1995). Comparison of pulmonary responses of 

asthmatic and nonasthmatic subjects performing light exercise while exposed to a low level of ozone. 

Toxicol Ind Health 11: 369-385. 

Horstman, DH; Folinsbee, LJ; Ives, PJ; Abdul-Salaam, S; McDonnell, WF. (1990). Ozone concentration and 

pulmonary response relationships for 6.6-hour exposures with five hours of moderate exercise to 0.08, 0.10, 

and 0.12 ppm. Am J Respir Crit Care Med 142: 1158-1163. 

Howden, Lindsay M. and Meyer, Julie A. (2011). U.S. Census Bureau, 2010 Census Briefs, C2010BR-03, Age and 

Sex Composition: 2010, U.S. Department of Commerce, Economics and Statistics Administration, U.S. 

Census Bureau, Washington, DC 20233. http://www.census.gov/prod/cen2010/briefs/c2010br-03.pdf 

Hurst, DJ; Gardner, DE; Coffin, DL. (1970). Effect of ozone on acid hydrolases of the pulmonary alveolar 

macrophage. J Reticuloendothel Soc 8: 288-300. 

Hwang, BF; Lee, YL; Lin, YC; Jaakkola, JJK; Guo, YL. (2005). Traffic related air pollution as a determinant of 

asthma among Taiwanese school children. Thorax 60: 467-473. 

Islam, T; Berhane, K; McConnell, R; Gauderman, WJ; Avol, E; Peters, JM; Gilliland, FD. (2009). Glutathione-Stransferase (GST) P1, GSTM1, exercise, ozone and asthma incidence in school children. Thorax 64: 197-

202. http://dx.doi.org/10.1136/thx.2008.099366 

Islam, T; McConnell, R; Gauderman, WJ; Avol, E; Peters, JM; Gilliland, FD. (2008). Ozone, oxidant defense genes 

and risk of asthma during adolescence. Am J Respir Crit Care Med 177: 388-395. 

http://dx.doi.org/10.1164/rccm.200706-863OC 

Islam, T; Gauderman, WJ; Berhane, K; McConnell, R; Avol, E; Peters, JM; Gilliland, FD. (2007). The relationship 

between air pollution, lung function and asthma in adolescents. Thorax 62: 957-963. 

Ito, K; Thurston, GD; Silverman, RA. (2007). Characterization of PM2.5, gaseous pollutants, and meteorological 

interactions in the context of time-series health effects models. J Expo Sci Environ Epidemiol 17: S45-S60. 

Jacquemin, B; Kauffmann, F; Pin, I; Le Moual, N; Bousquet, J; Gormand, F; Just, J; Nadif, R; Pison, C; Vervloet, 

D; Künzli, N; Siroux, V. (In Press) Air pollution and asthma control in the epidemiological study on the 

genetics and environment of asthma. J Epidemiol Community Health. 

http://dx.doi.org/10.1136/jech.2010.130229 

Jalaludin, BB; O'Toole, BI; Leeder, S. R. (2004). Acute effects of urban ambient air pollution on respiratory 

symptoms, asthma medication use, and doctor visits for asthma in a cohort of Australian children. Environ 

Res 95: 32-42. http://dx.doi.org/10.1016/S0013-9351 (03)00038-0 

Jerrett, M; Burnett, RT; Pope, CA, III; Ito, K; Thurston, G; Krewski, D; Shi, Y; Calle, E; Thun, M. (2009). Longterm ozone exposure and mortality. N Engl J Med 360: 1085-1095. 

http://dx.doi.org/10.1056/NEJMoa0803894 
3-143 

Joad, JP; Kott, KS; Bric, JM; Peake, JL; Plopper, CG; Schelegle, ES; Gershwin, LJ; Pinkerton, KE. (2006). 

Structural and functional localization of airway effects from episodic exposure of infant monkeys to 

allergen and/or ozone. Toxicol Appl Pharmacol 214: 237-243. 

Johnston, RA; Schwartzman, IN; Flynt, L; Shore, SA. (2005). Role of interleukin-6 in murine airway responses to 

ozone. Am J Physiol Lung Cell Mol Physiol 288: L390-L397 

Jorres, R; Nowak, D; Magnussen, H; Speckin, P; Koschyk, S. (1996). The effect of ozone exposure on allergen 

responsiveness in subjects with asthma or rhinitis. Am J Respir Crit Care Med 153: 56-64. 

Just, J; Segala, C; Sahraoui, F; Priol, G; Grimfeld, A; Neukirch, F. (2002). Short-term health effects of particulate 

and photochemical air pollution in asthmatic children. Eur Respir J 20: 899-906. 

http://dx.doi.org/10.1183/09031936.02.00236902 

Kan, H; London, SJ; Chen, G; Zhang, Y; Song, G; Zhao, N; Jiang, L; Chen, B. (2008). Season, sex, age, and 

education as modifiers of the effects of outdoor air pollution on daily mortality in Shanghai, China: The 

Public Health and Air Pollution in Asia (PAPA) Study. Environ Health Perspect 116: 1183-1188. 

http://dx.doi.org/10.1289/ehp.10851 

Katsouyanni, K; Samet, JM; Anderson, HR; Atkinson, R; Le Tertre, A; Medina, S; Samoli, E; Touloumi, G; 

Burnett, RT; Krewski, D; Ramsay, T; Dominici, F; Peng, RD; Schwartz, J; Zanobetti, A. (2009). Air 

pollution and health: A European and North American approach (APHENA). (Research Report 142). 

Boston, MA: Health Effects Institute. http://pubs.healtheffects.org/view.php?id=327 

Kehrl, HR; Peden, DB; Ball, BA; Folinsbee, LJ; Horstman, DH. (1999). Increased specific airway reactivity of 

persons with mild allergic asthma after 7.6 hours of exposure to 0.16 ppm ozone. J Allergy Clin Immunol 

104: 1198-1204. 

Khatri, SB; Holguin, FC; Ryan, PB; Mannino, D; Erzurum, SC; Teague, WG. (2009). Association of ambient ozone 

exposure with airway inflammation and allergy in adults with asthma. J Asthma 46: 777-785. 

http://dx.doi.org/10.1080/02770900902779284 

Kim, CS; Alexis, NE; Rappold, AG; Kehrl, H; Hazucha, MJ; Lay, JC; Schmitt, MT; Case, M; Devlin, RB; Peden, 

DB; Diaz-Sanchez, D. (2011). Lung function and inflammatory responses in healthy young adults exposed 

to 0.06 ppm ozone for 6.6 hours. Am J Respir Crit Care Med 183: 1215-1221. 

http://dx.doi.org/10.1164/rccm.201011-1813OC 

Koren, HS, RB Devlin, DE Graham; R Mann; MP Mcgee; DH Horstman; WJ Kozumbo; S Becker; DE House; WF 

McDonnell; PA Bromberg. (1989). Ozone-induced inflammation in the lower airways of human subjects. 

Am. Rev. Respir. Dis. 139:407-415. 

Korrick, SA; Neas, LM; Dockery, DW; Gold, DR; Allen, GA; Hill, LB; Kimball, KD; Rosner, BA; Speizer, FE. 

(1998). Effects of ozone and other pollutants on the pulmonary function of adult hikers. Environ Health 

Perspect 106: 93-99. http://dx.doi.org/10.1289/ehp.9810693 

Kreit, JW; Gross, KB; Moore, TB; Lorenzen, TJ; D'Arcy, J; Eschenbacher, WL. (1989). Ozone-induced changes in 

pulmonary function and bronchial responsiveness in asthmatics. J Appl Physiol 66: 217-222. 

Kulle, TJ; Sauder, LR; Kerr, HD; Farrell, BP; Bermel, MS; Smith, DM. (1982). Duration of pulmonary function 

adaptation to ozone in humans. Am Ind Hyg Assoc J 43: 832-837. 

Künzli, N. (2012). Is air pollution of the 20th century a cause of current asthma hospitalisations? [Editorial]. Thorax 

67: 2-3. http://dx.doi.org/10.1136/thoraxjnl-2011-200919 
3-144 

Larsen, ST; Matsubara, S; Mcconville, G; Poulsen, SS; Gelfand, EW. (2010). Ozone increases airway 

hyperreactivity and mucus hyperproduction in mice previously exposed to allergen. J Toxicol Environ 

Health A 73: 738-747. http://dx.doi.org/10.1080/15287391003614034 

Lee, YL; McConnell, R; Berhane, K; Gilliland, FD. (2009). Ambient ozone modifies the effect of tumor necrosis 

factor G-308A on bronchitic symptoms among children with asthma. Allergy 64: 1342-1348. 

http://dx.doi.org/10.1111/j.1398-9995.2009.02014.x 

Lewis, TC; Robins, TG; Dvonch, JT; Keeler, GJ; Yip, FY; Mentz, GB; Lin, X; Parker, EA; Israel, BA; Gonzalez, L; 

Hill, Y. (2005). Air pollution-associated changes in lung function among asthmatic children in Detroit. 

Environ Health Perspect 113: 1068-1075. 

Lin, S; Bell, EM; Liu, W; Walker, RJ; Kim, NK; Hwang, SA. (2008a). Ambient ozone concentration and hospital 

admissions due to childhood respiratory diseases in New York State, 1991-2001. Environ Res 108: 42-47. 

http://dx.doi.org/10.1016/j.envres.2008.06.007 

Lin, S; Liu, X; Le, LH; Hwang, SA. (2008b). Chronic exposure to ambient ozone and asthma hospital admissions 

among children. Environ Health Perspect 116: 1725-1730. http://dx.doi.org/10.1289/ehp.11184 

Linn, WS; Shamoo, DA; Anderson, KR; Peng, RC; Avol, EL; Hackney, JD; Gong, H, Jr. (1996). Short-term air 

pollution exposures and responses in Los Angeles area schoolchildren. J Expo Sci Environ Epidemiol 6: 

449-472. 

Liu, L; Poon, R; Chen, L; Frescura, AM; Montuschi, P; Ciabattoni, G; Wheeler, A; Dales, R. (2009). Acute effects 

of air pollution on pulmonary function, airway inflammation, and oxidative stress in asthmatic children. 

Environ Health Perspect 117: 668-674. http://dx.doi.org/10.1289/ehp11 

Mar, TF; Koenig, JQ. (2009). Relationship between visits to emergency departments for asthma and ozone exposure 

in greater Seattle, Washington. Ann Allergy Asthma Immunol 103: 474-479. 

McConnell, R; Berhane, K; Gilliland, F; London, SJ; Islam, T; Gauderman, WJ; Avol, E; Margolis, HG; Peters, JM. 

(2002). Asthma in exercising children exposed to ozone: A cohort study. Lancet 359: 386-391. 

McConnell, R; Islam, T; Shankardass, K; Jerrett, M; Lurmann, F; Gilliland, F; Gauderman, J; Avol, E; Kuenzli, N; 

Yao, L; Peters, J; Berhane, K. (2010). Childhood incident asthma and traffic-related air pollution at home 

and school. Environ Health Perspect 118: 1021-1026. 

McDonnell, WF. (1996). Individual variability in human lung function responses to ozone exposure. Environ 

Toxicol Pharmacol 2: 171-175. 

McDonnell, WF; Abbey, DE; Nishino, N; Lebowitz, MD. (1999). Long-term ambient ozone concentration and the 

incidence of asthma in nonsmoking adults: the Ahsmog study. Environ Res 80: 110-121. 

McDonnell, WF; Chapman, RS; Horstman, DH; Leigh, MW; Abdul-Salaam, S. (1985). A comparison of the 

responses of children and adults to acute ozone exposure. 

McDonnell, WF; Kehrl, HR; Abdul-Salaam, S; Ives, PJ; Folinsbee, LJ; Devlin, RB; O'Neil, JJ; Horstman, DH. 

(1991). Respiratory response of humans exposed to low levels of ozone for 6.6 hours. Arch Environ Occup 

Health 46: 145-150. 

McDonnell, WF; Stewart, PW; Smith, MV; Kim, CS; Schelegle, ES. (2012). Prediction of lung function response 

for populations exposed to a wide range of ozone conditions. Inhal Toxicol 24: 619-633. 

Medina-Ramón, M; Schwartz, J. (2008). Who is more vulnerable to die from ozone air pollution? Epidemiology 19: 

672-679. 
3-145 

Medina-Ramon, M; Zanobetti, A; Schwartz, J. (2006). The effect of ozone and PM10 on hospital admissions for 

pneumonia and chronic obstructive pulmonary disease: A national multicity study. Am J Epidemiol 163: 

579-588. http://dx.doi.org/10.1093/aje/kwj078 

Meng, YY; Rull, RP; Wilhelm, M; Lombardi, C; Balmes, J; Ritz, B. (2010). Outdoor air pollution and uncontrolled 

asthma in the San Joaquin Valley, California. J Epidemiol Community Health 64: 142-147. 

http://dx.doi.org/10.1136/jech.2008.083576 

Middleton, N; Yiallouros, P; Kleanthous, S; Kolokotroni, O; Schwartz, J; Dockery, DW; Demokritou, P; Koutrakis, 

P. (2008). A 10-year time-series analysis of respiratory and cardiovascular morbidity in Nicosia, Cyprus: 

The effect of short-term changes in air pollution and dust storms. Environ Health 7: 39. 

Moore, K; Neugebauer, R; Lurmann, F; Hall, J; Brajer, V; Alcorn, S; Tager, I. (2008). Ambient ozone 

concentrations cause increased hospitalizations for asthma in children: An 18-year study in Southern 

California. Environ Health Perspect 116: 1063-1070. http://dx.doi.org/10.1289/ehp.10497 

Mortimer, KM; Neas, LM; Dockery, DW; Redline, S; Tager, IB. (2002). The effect of air pollution on inner-city 

children with asthma. Eur Respir J 19: 699-705. http://dx.doi.org/10.1183/09031936.02.00247102 

Mudway, IS; Kelly, FJ. (2000). Ozone and the lung: A sensitive issue. Mol Aspects Med 21: 1-48. 

Mudway, IS; Stenfors, N; Blomberg, A; Helleday, R; Dunster, C; Marklund, SL; Frew, AJ; Sandstrom, T; Kelly, FJ. 

(2001). Differences in basal airway antioxidant concentrations are not predictive of individual 

responsiveness to ozone: A comparison of healthy and mild asthmatic subjects. Free Radic Biol Med 31: 

962-974. 

Mudway, IS; Kelly, FJ. (2004). An investigation of inhaled ozone dose and the magnitude of airway inflammation in 

healthy adults. Am J Respir Crit Care Med 169: 1089-1095. 

Neidell, M. (2009). Information, avoidance behavior, and health: The effect of ozone on asthma hospitalizations. 

Journal of Human Resources 44: 450-478. 

Neidell, M; Kinney, PL. (2010). Estimates of the association between ozone and asthma hospitalizations that 

account for behavioral responses to air quality information. Environ Sci Pol 13: 97-103. 

http://dx.doi.org/10.1016/j.envsci.2009.12.006 

NHLBI (National Institutes of Health, National Heart Lung and Blood Institute). (2007). Expert panel report 3: 

guidelines for the diagnosis and management of asthma. (07-4051). Bethesda, MD: National Institute of 

Health. 

NRC (2008). Estimating Mortality Risk Reduction and Economic Benefits from Controlling Ozone Air Pollution. 

Washington, DC: The National Academies Press. 

O'Connor, GT; Neas, L; Vaughn, B; Kattan, M; Mitchell, H; Crain, EF; III, ER; Gruchalla, R; Morgan, W; Stout, J; 

Adams, GK; Lippmann, M. (2008). Acute respiratory health effects of air pollution on children with asthma 

in U.S. inner cities. J Allergy Clin Immunol 121: 1133-1139. http://dx.doi.org/10.1016/j.jaci.2008.02.020 

O*Net OnLine (2012). Work Context — Outdoors, Exposed to Weather. 

http://www.onetonline.org/find/descriptor/result/4.C.2.a.1.c?a=1 

Ostro, B; Lipsett, M; Mann, J; Braxton-Owens, H; White, M. (2001). Air pollution and exacerbation of asthma in 

African-American children in Los Angeles. Epidemiology 12: 200-208. 

Peden, DB; Boehlecke, B; Horstman, D; Devlin, R. (1997). Prolonged acute exposure to 0.16 ppm ozone induces 

eosinophilic airway inflammation in asthmatic subjects with allergies. J Allergy Clin Immunol 100: 802-

808. 
3-146 

Peden, DB; Setzer, RW, Jr; Devlin, RB. (1995). Ozone exposure has both a priming effect on allergen-induced 

responses and an intrinsic inflammatory action in the nasal airways of perennially allergic asthmatics. Am J 

Respir Crit Care Med 151: 1336-1345. 

Pellegrino, R; Viegi, G; Brusasco, V; Crapo, RO; Burgos, F; Casaburi, R; Coates, A; van der Grinten, CP; 

Gustafsson, P; Hankinson, J; Jensen, R; Johnson, DC; MacIntyre, N; McKay, R; Miller, MR; Navajas, D; 

Pedersen, OF; Wanger, J. (2005). Interpretative strategies for lung function tests. Eur Respir J 26: 948-968. 

http://dx.doi.org/10.1183/09031936.05.00035205 

Plopper, CG; Smiley-Jewell, SM; Miller, LA; Fanucchi, MV; Evans, MJ; Buckpitt, AR; Avdalovic, M; Gershwin, 

LJ; Joad, JP; Kajekar, R; Larson, S; Pinkerton, KE; Van Winkle, LS; Schelegle, ES; Pieczarka, EM; Wu, 

R; Hyde, DM. (2007). Asthma/allergic airways disease: Does postnatal exposure to environmental 

toxicants promote airway pathobiology? Toxicol Pathol 35: 97-110. 

http://dx.doi.org/10.1080/01926230601132030 

Pope, CA, III; Burnett, RT; Thun, MJ; Calle, EE; Krewski, D; Ito, K; Thurston, GD. (2002). Lung cancer, 

cardiopulmonary mortality, and long-term exposure to fine particulate air pollution. JAMA 287: 1132-

1141. 

Qian, Z; Lin, HM; Chinchilli, VM; Lehman, EB; Duan, Y; Craig, TJ; Wilson, WE; Liao, D; Lazarus, SC; Bascom, 

R. (2009). Interaction of ambient air pollution with asthma medication on exhaled nitric oxide among 

asthmatics. Arch Environ Occup Health 64: 168-176. http://dx.doi.org/10.1080/19338240903240616 

Que, LG; Stiles, JV; Sundy, JS; Foster, WM. (2011). Pulmonary function, bronchial reactivity, and epithelial 

permeability are response phenotypes to ozone and develop differentially in healthy humans. J Appl 

Physiol 111: 679-687. http://dx.doi.org/10.1152/japplphysiol.00337.2011 

Rabinovitch, N; Zhang, LN; Murphy, JR; Vedal, S; Dutton, SJ; Gelfand, EW. (2004). Effects of wintertime ambient 

air pollutants on asthma exacerbations in urban minority children with moderate to severe disease. J 

Allergy Clin Immunol 114: 1131-1137. http://dx.doi.org/10.1016/j.jaci.2004.08.026 

Rage, E; Siroux, V; Kunzli, N; Pin, I; Kauffmann, F. (2009). Air pollution and asthma severity in adults. Occup 

Environ Med 66: 182-188. http://dx.doi.org/10.1136/oem.2007.038349 

Riediker, M; Monn, C; Koller, T; Stahel, WA; Wuthrich, B. (2001). Air pollutants enhance rhinoconjunctivitis 

symptoms in pollen-allergic individuals. Ann Allergy Asthma Immunol 87: 311-318. 

http://dx.doi.org/10.1016/S1081-1206 (10)62246-6 

Rojas-Martinez, R; Perez-Padilla, R; Olaiz-Fernandez, G; Mendoza-Alvarado, L; Moreno-Macias, H; Fortoul, T; 

Mcdonnell, W; Loomis, D; Romieu, I. (2007). Lung function growth in children with long-term exposure 

to air pollutants in Mexico City. Am J Respir Crit Care Med 176: 377-384. 

Romieu, I; Meneses, F; Ruiz, S; Huerta, J; Sienra, JJ; White, M; Etzel, R; Hernandez, M. (1997). Effects of 

intermittent ozone exposure on peak expiratory flow and respiratory symptoms among asthmatic children 

in Mexico City. Arch Environ Occup Health 52: 368-376. 

Romieu, I; Meneses, F; Ruiz, S; Sienra, JJ; Huerta, J; White, MC; Etzel, RA. (1996). Effects of air pollution of the 

respiratory health of asthmatic children living in Mexico City. Am J Respir Crit Care Med 154: 300-307. 

Romieu, I; Ramirez-Aguilar, M; Sienra-Monge, JJ; Moreno-Macias, H; Del Rio-Navarro, BE; David, G; Marzec, J; 

Hernandez-Avila, M; London, S. (2006). GSTM1 and GSTP1 and respiratory health in asthmatic children 

exposed to ozone. Eur Respir J 28: 953-959. http://dx.doi.org/10.1183/09031936.06.00114905 
3-147 

Romieu, I; Barraza-Villarreal, A; Escamilla-Núñez, C; Texcalac-Sangrador, JL; Hernandez-Cadena, L; DíazSánchez, D; De Batlle, J; Del Rio-Navarro, BE. (2009). Dietary intake, lung function and airway 

inflammation in Mexico City school children exposed to air pollutants. Respir Res 10: 122. 

Romieu, I; Barraza-Villarreal, A; Escamilla-Nunez, C; Almstrand, AC; Diaz-Sanchez, D; Sly, PD; Olin, AC. (2008). 

Exhaled breath malondialdehyde as a marker of effect of exposure to air pollution in children with asthma. 

J Allergy Clin Immunol 121: 903-909. http://dx.doi.org/10.1016/j.jaci.2007.12.004 

Romieu, I; Sienra-Monge, JJ; Ramirez-Aguilar, M; Tellez-Rojo, MM; Moreno-Macias, H; Reyes-Ruiz, NI; Del RioNavarro, BE; Ruiz-Navarro, MX; Hatch, G; Slade, R; Hernandez-Avila, M. (2002). Antioxidant 

supplementation and lung functions among children with asthma exposed to high levels of air pollutants. 

Am J Respir Crit Care Med 166: 703-709. 

Ross, MA; Persky, VW; Scheff, PA; Chung, J; Curtis, L; Ramakrishnan, V; Wadden, RA; Hryhorczuk, DO. (2002). 

Effect of ozone and aeroallergens on the respiratory health of asthmatics. Arch Environ Occup Health 57: 

568-578. http://dx.doi.org/10.1080/00039890209602090 

Salam, MT; Islam, T; Gauderman, WJ; Gilliland, FD. (2009). Roles of arginase variants, atopy, and ozone in 

childhood asthma. J Allergy Clin Immunol 123: 596-602. http://dx.doi.org/10.1016/j.jaci.2008.12.020 

Samet, J.M. (2011). Clean Air Scientific Advisory Committee (CASAC) Response to Charge Questions on the 

Reconsideration of the 2008 Ozone National Ambient Air Quality Standards. EPA-CASAC-11-004. 

March 30, 2011. Available online at: 

http://yosemite.epa.gov/sab/sabproduct.nsf/0/F08BEB48C1139E2A8525785E006909AC/$File/EPACASAC-11-004-unsigned+.pdf 

Samoli, E; Zanobetti, A; Schwartz, J; Atkinson, R; Le Tertre, A; Schindler, C; Pérez, L; Cadum, E; Pekkanen, J; 

Paldy, A; Touloumi, G; Katsouyanni, K. (2009). The temporal pattern of mortality responses to ambient 

ozone in the APHEA project. J Epidemiol Community Health 63: 960-966. 

http://dx.doi.org/10.1136/jech.2008.084012 

Sasser, E. (2014) Memo Responding to Request for Revised Ozone HREA Chapter 7 Appendix Tables. May 9, 

2014. Available at http://www.epa.gov/ttn/naaqs/standards/ozone/s_o3_2008_rea.html 

Scannell, C; Chen, L; Aris, RM; Tager, I; Christian, D; Ferrando, R; Welch, B; Kelly, T; Balmes, JR. (1996). 

Greater ozone-induced inflammatory responses in subjects with asthma. Am J Respir Crit Care Med 154: 

24-29. 

Schelegle, ES; Adams, WC; Walby, WF; Marion, MS. (2012). Modelling of individual subject ozone exposure 

response kinetics. Inhal Toxicol 24: 401-415. http://dx.doi.org/10.3109/08958378.2012.683891 

Schelegle, ES; Miller, LA; Gershwin, LJ; Fanucchi, MV; Van Winkle, LS; Gerriets, JE; Walby, WF; Mitchell, V; 

Tarkington, BK; Wong, VJ; Baker, GL; Pantle, LM; Joad, JP; Pinkerton, KE; Wu, R; Evans, MJ; Hyde, 

DM; Plopper, CG. (2003). Repeated episodes of ozone inhalation amplifies the effects of allergen 

sensitization and inhalation on airway immune and structural development in Rhesus monkeys. Toxicol 

Appl Pharmacol 191: 74-85. 

Schelegle, ES; Morales, CA; Walby, WF; Marion, S; Allen, RP. (2009). 6.6-hour inhalation of ozone concentrations 

from 60 to 87 parts per billion in healthy humans. Am J Respir Crit Care Med 180: 265-272. 

http://dx.doi.org/10.1164/rccm.200809-1484OC 

Schelegle, ES; Siefkin, AD; McDonald, RJ. (1991). Time course of ozone-induced neutrophilia in normal humans. 

Am J Respir Crit Care Med 143: 1353-1358. 
3-148 

Schildcrout, JS; Sheppard, L; Lumley, T; Slaughter, JC; Koenig, JQ; Shapiro, GG. (2006). Ambient air pollution 

and asthma exacerbations in children: An eight-city analysis. Am J Epidemiol 164: 505-517. 

http://dx.doi.org/10.1093/aje/kwj225 

Schiller JS, Lucas JW, Ward BW, Peregoy JA. (2012). Summary health statistics for U.S. adults: National Health 

Interview Survey, 2010. National Center for Health Statistics. Vital Health Stat 10(252). 

http://www.cdc.gov/nchs/data/series/sr_10/sr10_252.pdf 

Sienra-Monge, JJ; Ramirez-Aguilar, M; Moreno-Macias, H; Reyes-Ruiz, NI; Del Rio-Navarro, BE; Ruiz-Navarro, 

MX; Hatch, G; Crissman, K; Slade, R; Devlin, RB; Romieu, I. (2004). Antioxidant supplementation and 

nasal inflammatory responses among young asthmatics exposed to high levels of ozone. Clin Exp Immunol 

138: 317-322. http://dx.doi.org/10.1111/j.1365-2249.2004.02606.x 

Silverman, RA; Ito, K. (2010). Age-related association of fine particles and ozone with severe acute asthma in New 

York City. J Allergy Clin Immunol 125: 367-373. http://dx.doi.org/10.1016/j.jaci.2009.10.061 

Smith, RL; Xu, B; Switzer, P. (2009). Reassessing the relationship between ozone and short-term mortality in U.S. 

urban communities. Inhal Toxicol 21: 37-61. http://dx.doi.org/10.1080/08958370903161612 

Spektor, DM; Lippmann, M; Lioy, PJ; Thurston, GD; Citak, K; James, DJ; Bock, N; Speizer, FE; Hayes, C. 

(1988a). Effects of ambient ozone on respiratory function in active, normal children. Am Rev Respir Dis 

137: 313-320. 

Spektor, DM; Lippmann, M; Thurston, GD; Lioy, PJ; Stecko, J; O'Connor, G; Garshick, E; Speizer, FE; Hayes, C. 

(1988b). Effects of ambient ozone on respiratory function in healthy adults exercising outdoors. Am Rev 

Respir Dis 138: 821-828. 

SSDAN CensusScope (Social Science Data Analysis Network, CensusScope). (2010). United States: Age 

distribution [Database]. Ann Arbor, Michigan: Social Science Data Analysis Network. Retrieved from 

http://www.censusscope.org/us/chart_age.html 

Stafoggia, M; Forastiere, F; Faustini, A; Biggeri, A; Bisanti, L; Cadum, E; Cernigliaro, A; Mallone, S; Pandolfi, P; 

Serinelli, M; Tessari, R; Vigotti, MA; Perucci, CA. (2010). Susceptibility factors to ozone-related 

mortality: A population-based case-crossover analysis. Am J Respir Crit Care Med 182: 376-384. 

http://dx.doi.org/10.1164/rccm.200908-1269OC 

Stenfors, N; Pourazar, J; Blomberg, A; Krishna, MT; Mudway, I; Helleday, R; Kelly, FJ; Frew, AJ; Sandstrom, T. 

(2002). Effect of ozone on bronchial mucosal inflammation in asthmatic and healthy subjects. Respir Med 

96: 352-358. 

Stieb, DM; Szyszkowicz, M; Rowe, BH; Leech, JA. (2009). Air pollution and emergency department visits for 

cardiac and respiratory conditions: A multi-city time-series analysis. Environ Health Global Access Sci 

Source 8: 25. http://dx.doi.org/10.1186/1476-069X-8-25 

Strickland, MJ; Darrow, LA; Klein, M; Flanders, WD; Sarnat, JA; Waller, LA; Sarnat, SE; Mulholland, JA; Tolbert, 

PE. (2010). Short-term associations between ambient air pollutants and pediatric asthma emergency 

department visits. Am J Respir Crit Care Med 182: 307-316. http://dx.doi.org/10.1164/rccm.200908-

1201OC 

Stylianou, M; Nicolich, MJ. (2009). Cumulative effects and threshold levels in air pollution mortality: Data analysis 

of nine large US cities using the NMMAPS dataset. Environ Pollut 157: 2216-2223. 

http://dx.doi.org/10.1016/j.envpol.2009.04.011 
3-149 

Thaller, EI; Petronella, SA; Hochman, D; Howard, S; Chhikara, RS; Brooks, EG. (2008). Moderate increases in 

ambient PM2.5 and ozone are associated with lung function decreases in beach lifeguards. J Occup Environ 

Med 50: 202-211. http://dx.doi.org/10.1097/JOM.0b013e31816386b4 

Thurston, GD; Lippmann, M; Scott, MB; Fine, JM. (1997). Summertime haze air pollution and children with 

asthma. Am J Respir Crit Care Med 155: 654-660. 

Tolbert, PE; Klein, M; Peel, JL; Sarnat, SE; Sarnat, JA. (2007). Multipollutant modeling issues in a study of ambient 

air quality and emergency department visits in Atlanta. J Expo Sci Environ Epidemiol 17: S29-S35. 

http://dx.doi.org/10.1038/sj.jes.7500625 

Torres, A; Utell, MJ; Morow, PE; Voter, KZ; Whitin, JC; Cox, C; Looney, RJ; Speers, DM; Tsai, Y; Frampton, 

MW. (1997). Airway inflammation in smokers and nonsmokers with varying responsiveness to ozone. Am 

J Respir Crit Care Med 156: 728-736. 

Trenga, CA; Koenig, JQ; Williams, PV. (2001). Dietary antioxidants and ozone-induced bronchial 

hyperresponsiveness in adults with asthma. Arch Environ Occup Health 56: 242-249. 

U.S. EPA (1996). Air quality criteria for ozone and related photochemical oxidants [EPA Report]. (EPA/600/P93/004aF, cF). U.S. Environmental Protection Agency, Research Triangle Park, NC. 

U.S. EPA (2006). Air Quality Criteria for Ozone and Related Photochemical Oxidants (2006 Final). U.S. 

Environmental Protection Agency, Washington, DC. EPA/600/R-05/004aF-cF. March 2006. Available at: 

http://www.epa.gov/ttn/naaqs/standards/ozone/s_o3_cr_cd.html 

U.S. EPA (2007). Review of the National Ambient Air Quality Standards for Ozone: Policy Assessment of 

Scientific and Technical Information, OAQPS Staff Paper. EPA-452/R-07-007 

U.S. EPA (2009). Integrated Science Assessment for Particulate Matter (Final Report). U.S. Environmental 

Protection Agency, Washington, DC, EPA/600/R-08/139F, 2009. 

U.S. EPA (2013). Integrated Science Assessment of Ozone and Related Photochemical Oxidants (Final Report). 

U.S. Environmental Protection Agency, Washington, DC. EPA/600/R-10/076F. Available at: 

http://www.epa.gov/ttn/naaqs/standards/ozone/s_o3_2008_isa.html 

U.S. EPA (2014). Health Risk and Exposure Assessment for Ozone. Office of Air Quality Planning and Standards, 

Research Triangle Park, NC, 27711. EPA-452/R-14-004a. August 2014. Available at: 

http://www.epa.gov/ttn/naaqs/standards/ozone/s_o3_index.html 

Ulmer, C; Kopp, M; Ihorst, G; Frischer, T; Forster, J; Kuehr, J. (1997). Effects of ambient ozone exposures during 

the spring and summer of 1994 on pulmonary function of schoolchildren. Pediatr Pulmonol 23: 344-353. 

http://dx.doi.org/10.1002/(SICI)1099-0496(199705)23:5<344::AID-PPUL6>3.0.CO;2-K 

Vagaggini, B; Cianchetti, S; Bartoli, M; Ricci, M; Bacci, E; Dente, FL; Di Franco, A; Paggiaro, P. (2007). 

Prednisone blunts airway neutrophilic inflammatory response due to ozone exposure in asthmatic subjects. 

Respiration 74: 61-58. http://dx.doi.org/10.1159/000096078 

Vagaggini, B; Taccola, M; Clanchetti, S; Carnevali, S; Bartoli, ML; Bacci, E; Dente, FL; Di Franco, A; Giannini, D; 

Paggiaro, PL. (2002). Ozone exposure increases eosinophilic airway response induced by previous allergen 

challenge. Am J Respir Crit Care Med 166: 1073-1077. 

Vagaggini, B; Taccola, M; Conti, I; Carnevali, S; Cianchetti, S; Bartoli, ML; Bacci, E; Dente, FL; Di Franco, A; 

Giannini, D; Paggiaro, PL. (2001). Budesonide reduces neutrophilic but not functional airway response to 

ozone in mild asthmatics. Am J Respir Crit Care Med 164: 2172-2176. 
3-150 

Vedal, S.; Brauer, M.; White, R.; Petkau, J. (2003) Air pollution and daily mortality in a city with low levels of 

pollution. Environ. Health Perspect. 111: 45–51. 

Villeneuve, PJ; Chen, L; Rowe, BH; Coates, F. (2007). Outdoor air pollution and emergency department visits for 

asthma among children and adults: A case-crossover study in northern Alberta, Canada. Environ Health 

Global Access Sci Source 6: 40. http://dx.doi.org/10.1186/1476-069X-6-40 

Ward, DJ; Roberts, KT; Jones, N; Harrison, RM; Ayres, JG; Hussain, S; Walters, S. (2002). Effects of daily 

variation in outdoor particulates and ambient acid species in normal and asthmatic children. Thorax 57: 

489-502. http://dx.doi.org/10.1136/thorax.57.6.489 

Wenten, M; Gauderman, WJ; Berhane, K; Lin, PC; Peters, J; Gilliland, FD. (2009). Functional variants in the 

catalase and myeloperoxidase genes, ambient air pollution, and respiratory-related school absences: An 

example of epistasis in gene-environment interactions. Am J Epidemiol 170: 1494-1501. 

http://dx.doi.org/10.1093/aje/kwp310 

Wolff, G.T. (1995) Letter to EPA Administrator Carol Browner: “CASAC Closure on the Primary Standard Portion 

of the Staff Paper for Ozone” EPA-SAB-CASAC-LTR-96-002, November 30, 1995. 

Wong, CM; Vichit-Vadakan, N; Vajanapoom, N; Ostro, B; Thach, TQ; Chau, PY; Chan, EK; Chung, RY; Ou, CQ; 

Yang, L; Peiris, JS; Thomas, GN; Lam, TH; Wong, TW; Hedley, AJ; Kan, H; Chen, B; Zhao, N; London, 

SJ; Song, G; Chen, G; Zhang, Y; Jiang, L; Qian, Z; He, Q; Lin, HM; Kong, L; Zhou, D; Liang, S; Zhu, Z; 

Liao, D; Liu, W; Bentley, CM; Dan, J; Wang, B; Yang, N; Xu, S; Gong, J; Wei, H; Sun, H; Qin, Z. (2010). 

Part 5. Public health and air pollution in Asia (PAPA): A combined analysis of four studies of air pollution 

and mortality. In Public Health and Air Pollution in Asia (PAPA): Coordinated Studies of Short-Term 

Exposure to Air Pollution and Daily Mortality in Four Cities (pp. 377-418). Boston, MA: Health Effects 

Institute. http://pubs.healtheffects.org/view.php?id=348 

Wong, CM; Yang, L; Thach, TQ; Chau, PY; Chan, KP; Thomas, GN; Lam, TH; Wong, TW; Hedley, AJ; Peiris, JS. 

(2009). Modification by influenza on health effects of air pollution in Hong Kong. Environ Health Perspect 

117: 248-253. http://dx.doi.org/10.1289/ehp.11605 

Woodwell, D. A.; Cherry, D. K. (2004) National Ambulatory Medical Care Survey: 2002 summary. Hyattsville, 

MD: National Center for Health Statistics; DHHS publication no. (PHS) 2004-1250. (Advance data from 

vital and health statistics; no. 346). Available: http://www.cdc.gov/nchs/data/ad/ad346.pdf [3 August, 

2005] 

Xia, Y; Tong, H. (2006). Cumulative effects of air pollution on public health. Stat Med 25: 3548-3559. 

http://dx.doi.org/10.1002/sim.2446 

Yang, Q; Chen, Y; Krewski, D; Burnett, RT; Shi, Y; Mcgrail, KM. (2005). Effect of short-term exposure to low 

levels of gaseous pollutants on chronic obstructive pulmonary disease hospitalizations. Environ Res 99: 99-

105. http://dx.doi.org/10.1016/j.envres.2004.09.014 

Zanobetti, A; Schwartz, J. (2011). Ozone and survival in four cohorts with potentially predisposing diseases. Am J 

Respir Crit Care Med 184: 836-841. http://dx.doi.org/10.1164/rccm.201102-0227OC 

Zanobetti, A; Schwartz, J. (2008a). Is there adaptation in the ozone mortality relationship: A multi-city casecrossover analysis. Environ Health 7: 22. http://dx.doi.org/10.1186/1476-069X-7-22 

Zanobetti, A; Schwartz, J. (2008b). Mortality displacement in the association of ozone with mortality: An analysis 

of 48 cities in the United States. Am J Respir Crit Care Med 177: 184-189. 

http://dx.doi.org/10.1164/rccm.200706-823OC 
3-151 

Zanobetti, A; Schwartz, J. (2006). Air pollution and emergency admissions in Boston, MA. J Epidemiol Community 

Health 60: 890-895. http://dx.doi.org/10.1136/jech.2005.039834 
4-1 

4 CONSIDERATION OF ALTERNATIVE PRIMARY STANDARDS 

Having reached the conclusion that the currently available scientific evidence and 

exposure/risk information calls into question the adequacy of the current O3 standard, we next 

consider the following overarching question: 

 What is the range of potential alternative standards that are supported by the 

currently available scientific evidence and exposure/risk information, as reflected 

in the ISA and HREA respectively? 

To address this overarching question, in the sections below we evaluate a series of more specific 

questions related to the major elements of the NAAQS: indicator (section 4.1), averaging time 

(section 4.2), form (section 4.3), and level (section 4.4). In addressing these questions, we 

consider the currently available scientific evidence and exposure/risk information, including the 

evidence and information available at the time of the last review and that newly available in the 

current review, as assessed in the ISA and the HREA. In so doing, we note that the final decision 

by the Administrator in this review will consider these elements collectively in evaluating the 

health protection afforded by the primary standard.1

 

4.1 INDICATOR 

In the last review, EPA focused on O3 as the most appropriate indicator for a standard 

meant to provide protection against ambient photochemical oxidants. In this review, while the 

complex atmospheric chemistry in which O3 plays a key role has been highlighted, no 

alternatives to O3 have been advanced as being a more appropriate indicator for ambient 

photochemical oxidants. More specifically, the ISA noted that O3 is the only photochemical 

oxidant (other than NO2) that is routinely monitored and for which a comprehensive database 

exists (ISA section 3.6). Data for other photochemical oxidants (e.g., PAN, H2O2, etc.) typically 

have been obtained only as part of special field studies. Consequently, no data on nationwide 

patterns of occurrence are available for these other oxidants; nor are extensive data available on 

the relationships of concentrations and patterns of these oxidants to those of O3 (U.S. EPA, 2013, 

section 3.6). In its review of the second draft PA, CASAC concurred, stating “The indicator of 

ozone is appropriate based on its causal or likely causal associations with multiple adverse health 

outcomes and its representation of a class of pollutants known as photochemical oxidants” (Frey, 

2014, p. ii). 

 

1

We also take note of the 1997 review (discussed in section 1.3.1.2.3), in which O3 background concentrations were 

an additional consideration in EPA’s selection of a standard from among a range of scientifically acceptable 

alternatives. Background O3 is discussed in more detail in chapter 2 of this PA. 
4-2 

We further note that meeting an O3 standard can be expected to provide some degree of 

protection against potential health effects that may be independently associated with other 

photochemical oxidants, even though such effects are not discernible from currently available 

studies indexed by O3 alone. That is, since the precursor emissions that lead to the formation of 

O3 generally also lead to the formation of other photochemical oxidants, measures leading to 

reductions in population exposures to O3 can generally be expected to lead to reductions in 

population exposures to other photochemical oxidants. Taken together, we conclude that O3

remains the most appropriate indicator for a standard meant to provide protection against 

photochemical oxidants.2

4.2 AVERAGING TIME 

The EPA established the current 8-hour averaging time3 for the primary O3 NAAQS in 

1997 (62 FR 38856). The decision on averaging time in that review was based on numerous 

controlled human exposure and epidemiologic studies reporting associations between 6 to 8 hour 

O3 concentrations and adverse respiratory effects (62 FR 38861). It was also noted that a 

standard with a max 8-hour averaging time is likely to provide substantial protection against 

respiratory effects associated with 1-hour peak O3 concentrations. Similar conclusions were 

reached in the last O3 NAAQS review and thus, the 8-hour averaging time was retained in 2008. 

In the current review, we first consider the following question related to averaging time: 

 To what extent does the available evidence continue to support the 

appropriateness of a standard with an 8-hour averaging time? 

In reaching conclusions related to this question, staff considers causality judgments from the 

ISA, as well as results from the specific controlled human exposure and epidemiologic studies 

that informed those judgments. These considerations are described below in more detail. 

As an initial consideration with respect to the most appropriate averaging time for the O3

NAAQS, we note that the strongest evidence for O3-associated health effects is for respiratory 

effects following short-term exposures. More specifically, the ISA concludes that evidence 

relating short-term O3 exposures to respiratory effects is “sufficient to infer a causal 

relationship.” The ISA also judges that the evidence for short-term exposures to O3 indicates 

“likely to be” causal relationships with both cardiovascular effects and mortality (U.S. EPA, 

2013, section 2.5.2). Therefore, as in past reviews, the strength of the available scientific 

 

2

The D.C. Circuit upheld the use of O3 as the indicator for photochemical oxidants based on these same 

considerations. American Petroleum Inst. v. Costle, 665 F. 2d 1176, 1186 (D.C. Cir. 1981). 

3

This 8-hour averaging time reflects daily max 8-hour average O3 concentrations. 
4-3 

evidence provides strong support for a standard that protects the public health against short-term 

exposures to O3. 

In first considering the level of support available for specific short-term averaging times, 

we note the evidence available from controlled human exposure studies. As discussed in more 

detail in chapter 3 of this PA, substantial health effects evidence from controlled human 

exposure studies demonstrates that a wide range of respiratory effects (e.g., pulmonary function 

decrements, increases in respiratory symptoms, lung inflammation, lung permeability, decreased 

lung host defense, and airway hyperresponsiveness) occur in healthy adults following 6.6 hour 

exposures to O3 (EPA 2013, section 6.2.1.1). Compared to shorter exposure durations (e.g., 1-

hour), studies evaluating 6.6 hour exposures in healthy adults have reported respiratory effects at 

lower O3 exposure concentrations and at more moderate levels of exertion. 

We also note the strength of evidence from epidemiologic studies that have evaluated a 

wide variety of populations (e.g., including at-risk lifestages and populations, such as children 

and people with asthma, respectively). A number of different averaging times are used in O3

epidemiologic studies, with the most common being the max 1-hour concentration within a 24-

hour period (1-hour max), the max 8-hour average concentration within a 24-hour period (8-hour 

max), and the 24-hour average. These studies are discussed in chapter 3 of this PA, and are 

assessed in detail in chapter 6 of the ISA (U.S. EPA, 2013). Limited evidence from time-series 

and panel epidemiologic studies comparing risk estimates across averaging times does not 

indicate that one exposure metric is more consistently or strongly associated with respiratory 

health effects or mortality, though the ISA notes some evidence for “smaller O3 risk estimates 

when using a 24-hour average exposure metric” (EPA 2013, section 2.5.4.2; p. 2-31). For singleand multi-day average O3 concentrations, lung function decrements were associated with 1-hour 

max, 8-hour max, and 24-hour average ambient O3 concentrations, with no strong difference in 

the consistency or magnitude of association among the averaging times (EPA 2013, p. 6-71). 

Similarly, in studies of short-term exposure to O3 and mortality, Smith et al. (2009) and Darrow 

et al. (2011) have reported high correlations between risk estimates calculated using 24-hour 

average, 8-hour max, and 1-hour max averaging times (EPA 2013, p. 6-253). Thus, the 

epidemiologic evidence alone does not provide a strong basis for distinguishing between the 

appropriateness of 1-hour, 8-hour, and 24-hour averaging times. 

Considering the health information discussed above, we conclude that an 8-hour 

averaging time remains appropriate for addressing health effects associated with short-term 

exposures to ambient O3. An 8-hour averaging time is similar to the exposure periods evaluated 

in controlled human exposure studies, including recent studies that provide evidence for 

respiratory effects following exposures to O3 concentrations below the level of the current 

standard. In addition, epidemiologic studies provide evidence for health effect associations with 
4-4 

8-hour O3 concentrations, as well as with 1-hour and 24-hour concentrations. As in previous 

reviews, we note that a standard with an 8-hour averaging time (combined with an appropriate 

standard form and level) would also be expected to provide substantial protection against health 

effects attributable to 1-hour and 24-hour exposures (e.g., 62 FR 38861, July 18, 1997). In its 

review of the second draft PA, CASAC concurred stating that “the current 8-hour averaging time 

is justified by the combined evidence from epidemiologic and clinical studies” (Frey, 2014, p. 6).

The ISA also concludes that the evidence for long-term O3 exposures indicates that there 

is “likely to be a causal relationship” with respiratory effects (US EPA, 2013, chapter 7). Thus, 

in this review we also consider the extent to which currently available evidence and 

exposure/risk information suggests that a standard with an 8-hour averaging time can provide 

protection against respiratory effects associated with longer term exposures to ambient O3. In 

doing so, staff considers the following question: 

 To what extent does the available evidence and exposure/risk information indicate 

that a standard with the current 8-hour averaging time could provide protection 

against long-term exposures to ambient O3? 

In considering this issue in the last review of the O3 NAAQS, staff noted that “because long-term 

air quality patterns would be improved in areas coming into attainment with an 8-hr standard, the 

potential risk of health effects associated with long-term exposures would be reduced in any area 

meeting an 8-hr standard” (U.S. EPA, 2007, p. 6-57). 

 In the current review, we further evaluate this issue, with a focus on the “long-term” O3

metrics reported to be associated with mortality or morbidity in recent epidemiologic studies. As 

discussed in section 3.1.3, much of the recent evidence for such associations is based on studies 

that defined long-term O3 in terms of seasonal averages of daily max concentrations (e.g., 

seasonal averages of 1-hour or 8-hour daily max concentrations). 

As an initial consideration, we note the risk results from the HREA for respiratory 

mortality associated with long-term O3 concentrations. As discussed in section 3.2.3.2, HREA 

analyses indicate that as air quality is adjusted to just meet the current 8-hour standard, most 

urban case study areas are estimated to experience reductions in respiratory mortality associated 

with long-term O3 concentrations based on the seasonal averages of 1-hour daily max O3

concentrations evaluated in the study by Jerrett et al. (2009) (U.S. EPA, 2014, chapter 7). As air 

quality is adjusted to meet lower potential alternative standard levels, for standards based on 3-

year averages of the annual fourth-highest daily max 8-hour O3 concentrations, respiratory 

mortality risks are estimated to be reduced further in urban case study areas (section 

4.4.2.3,below). This analysis indicates that an O3 standard with an 8-hour averaging time, when 

coupled with an appropriate form and level, can reduce respiratory mortality reported to be 

associated with “long-term” O3 concentrations. 
4-5 

In further considering the study by Jerrett et al. (2009), we compare long-term O3

concentrations following air quality adjustment in urban case study areas (i.e., adjusted to meet 

the current and potential alternative 8-hour standards) to the concentrations present in study 

cities that provided the basis for the positive and statistically significant association with 

respiratory mortality. As indicated below (Table 4-3), this comparison suggests that a standard 

with an 8-hour averaging time can decrease seasonal averages of 1-hour daily max O3

concentrations, and can maintain those O3 concentrations below the seasonal average where we 

have the most confidence in the reported concentration-response relationship with respiratory 

mortality (see section 4.4.1 for further discussion). 

The HREA also conducted analyses evaluating the impacts of reducing regional NOX

emissions on the seasonal averages of 8-hour daily max O3 concentrations.4

 Seasonal averages of 

8-hour daily max O3 concentrations reflect long-term metrics that have been reported to be 

associated with respiratory morbidity effects in several recent O3 epidemiologic studies (e.g., 

Islam et al., 2008; Lin et al., 2008; Salam et al., 2009). The HREA analyses indicate that the 

large majority of the U.S. population lives in locations where reducing NOX emissions would be 

expected to result in decreases in seasonal averages of daily max 8-hour ambient O3

concentrations (U.S. EPA, 2014, section 8.2.3.2). Thus, consistent with the respiratory mortality 

risk estimates noted above, this analysis suggests that reductions in O3 precursor emissions in 

order to meet a standard with an 8-hour averaging time would also be expected to reduce the 

long-term O3 concentrations that have been reported in recent epidemiologic studies to be 

associated with respiratory morbidity. 

Taken together, we conclude that a standard with an 8-hour averaging time, coupled with 

the current 4th high form and an appropriate level, would be expected to provide appropriate 

protection against the long-term O3 concentrations that have been reported to be associated with 

respiratory morbidity and mortality. In its review of the second draft PA, CASAC concurred, 

stating that “The 8-hour averaging window also provides protection against the adverse impacts 

of long-term ozone exposures, which were found to be “likely causal” for respiratory effects and 

premature mortality” (Frey, 2014, p. 6). This issue is considered further, within the context of 

specific potential alternative standard levels, in section 4.4 below. 

4.3 FORM 

The “form” of a standard defines the air quality statistic that is to be compared to the 

level of the standard in determining whether an area attains the standard. The foremost 

consideration in selecting a form for potential alternative primary standards is the adequacy of 

 

4

Analyses are based on regional NOX reductions, which are effective in bringing down peak ambient O3

concentrations, but can have variable impacts on seasonal mean concentrations. 
4-6 

the public health protection provided by the combination of the form and the other elements of 

the standard. As such, in reaching staff conclusions regarding the appropriate form(s) to consider 

for a potential alternative primary O3 standard, we consider the following question: 

 To what extent do the available evidence and/or information continue to support 

the appropriateness of a standard with a form defined by the 3-year average of 

annual 4th-highest 8-hour daily max O3 concentrations? 

The EPA established the current form of the primary O3 NAAQS in 1997 (62 FR 38856). 

Prior to that time, the standard had a “1-expected-exceedance” form.5 An advantage of the 

current concentration-based form recognized in the 1997 review is that such a form better 

reflects the continuum of health effects associated with increasing ambient O3 concentrations. 

Unlike an expected exceedance form, a concentration-based form gives proportionally more 

weight to years when 8-hour O3 concentrations are well above the level of the standard than to 

years when 8-hour O3 concentrations are just above the level of the standard. It was judged 

appropriate to give more weight to higher O3 concentrations, given that available health evidence 

indicated a continuum of effects associated with exposures to varying concentrations of O3, and 

given that the extent to which public health is affected by exposure to ambient O3 is related to the 

actual magnitude of the O3 concentration, not just whether the concentration is above a specified 

level. 

During the 1997 review, EPA considered a range of alternative “concentration-based” 

forms, including the second-, third-, fourth- and fifth-highest daily max 8-hour concentrations in 

an O3 season. The fourth-highest daily max was selected, recognizing that a less restrictive form 

(e.g., fifth highest) would allow a relatively large percentage of sites to experience O3 peaks well 

above the level of the standard, and would allow more days on which the level of the standard 

may be exceeded when attaining the standard (62 FR 38856). Consideration was also given to 

setting a standard with a form that would provide a margin of safety against possible but 

uncertain chronic effects, and would provide greater stability to ongoing control programs.6

 A 

more restrictive form was not selected, recognizing that the differences in the degree of 

protection afforded by the alternatives were not well enough understood to use any such 

differences as a basis for choosing the most restrictive forms (62 FR 38856). 

In the 2008 review, EPA additionally considered the potential value of a percentile-based 

form. In doing so, EPA recognized that such a statistic is useful for comparing datasets of 

 

5

For a standard with a 1-expected-exceedance form to be met at an air quality monitoring site, the fourth-highest air 

quality value in 3 years, given adjustments for missing data, must be less than or equal to the level of the standard. 6

 See American Trucking Assn’s v. EPA, 283 F. 3d 355, 374-75 (D.C. Cir. 2002) (less stable implementation 

programs may be less effective, and therefore EPA can consider programmatic stability in determining the form of a 

NAAQS). 
4-7 

varying length because it samples approximately the same place in the distribution of air quality 

values, whether the dataset is several months or several years long. However, EPA concluded 

that a percentile-based statistic would not be effective in ensuring the same degree of public 

health protection across the country. Specifically, a percentile-based form would allow more 

days with higher air quality values in locations with longer O3 seasons relative to places with 

shorter O3 seasons. Thus, in the 2008 review EPA concluded that a form based on the nth-highest 

max O3 concentration would more effectively ensure that people who live in areas with different 

length O3 seasons receive the same degree of public health protection. 

Based on analyses for forms specified in terms of an nth-highest concentration (n ranged 

from 3 to 5), advice from CASAC, and public comment,7 the Administrator concluded that a 4thhighest daily max should be retained (73 FR 16465). In reaching this decision, the Administrator 

recognized that “there is not a clear health-based threshold for selecting a particular nth-highest 

daily maximum form of the standard” and that “the adequacy of the public health protection 

provided by the combination of the level and form is a foremost consideration” (73 FR 16475). 

Based on this, the Administrator judged that the existing form (4th-highest daily maximum 8-

hour average concentration) should be retained, recognizing the increase in public health 

protection provided by combining this form with a lower standard level (i.e., 75 ppb). 

The Administrator also recognized that it is important to have a form that provides 

stability with regard to implementation of the standard. In the case of O3, for example, he noted 

the importance of a form insulated from the impacts of the meteorological events that are 

conducive to O3 formation. Such events could have the effect of reducing public health 

protection, to the extent they result in frequent shifts in and out of attainment due to 

meteorological conditions. The Administrator noted that such frequent shifting could disrupt an 

area’s ongoing implementation plans and associated control programs (73 FR 16474). In his final 

decision, the Administrator judged that a “4th high form provides a stable target for implementing 

programs to improve air quality” (73 FR 16475). 

In the current review, we consider the extent to which newly available information 

provides support for consideration of alternative forms. In so doing, we take note of the 

conclusions of prior reviews summarized above. We recognize the value of an nth-high statistic 

over that of an expected exceedance or percentile-based form in the case of the O3 standard, for 

 

7

In the 2008 review, one group of commenters expressed the view that the standard was not adequate and supported 

a more health-protective form (e.g., a second- or third-highest daily max form). Another group of commenters 

expressed the view that the standard was adequate and did not provide any views on alternative forms that would be 

appropriate should the Administrator consider revisions to the standard. The Administrator considered the protection 

afforded by the combination of level and form in revising the standard in 2008 to 75 ppb, as a 3-year average of the 

annual fourth-highest daily max 8-hour concentrations (73 FR 16475). 
4-8 

the reasons summarized above. We additionally take note of the importance of stability in 

implementation to achieving the level of protection specified by the NAAQS. Specifically, we 

note that to the extent that areas engaged in implementing the O3 NAAQS frequently shift from 

meeting to violating the standard, it is possible that ongoing implementation plans and associated 

control programs could be disrupted, thereby reducing public health protection. 

In light of this, while giving foremost consideration to the adequacy of public health 

protection provided by the combination of all elements of the standard, including the form, we 

consider particularly findings from prior reviews with regard to the use of the nth-high metric. 

As noted above, the 4th-highest daily max was selected in recognition of the public health 

protection provided by this form, when coupled with an appropriate averaging time and level, 

and recognizing that such a form can provide stability for implementation programs. The 

currently available evidence and information does not call into question these conclusions from 

previous reviews. Moreover, in its review of the second draft PA, CASAC concurred that the O3

standard should be based on the fourth highest, daily maximum 8-hour average value (averaged 

over three years), stating that this form “provides health protection while allowing for atypical 

meteorological conditions that can lead to abnormally high ambient ozone concentrations which, 

in turn, provides programmatic stability” (Frey, 2014, p. 6). Thus a standard with the current 4th

high form, coupled with a level lower than 75 ppb as discussed below, would be expected to 

increase public health protection relative to the current standard while continuing to provide 

stability for implementation programs. Therefore, we conclude that it would be appropriate to 

consider retaining the current 4th-highest daily max form for an O3 standard with an 8-hour 

averaging time and a revised level, as discussed below. 

4.4 LEVEL 

In considering potential alternative standards levels to provide greater protection than that 

afforded by the current standard against O3-related adverse health effects, we address the 

following overarching question. 

 For an O3 standard defined in terms of the current indicator, averaging time, and 

form, what alternative levels are appropriate to consider in order to provide 

adequate public health protection against short- and long- term exposures to O3

in ambient air? 

In considering this question, we take into account the experimental and epidemiologic evidence 

as presented in the ISA, as well as the uncertainties and limitations associated with this evidence 

(section 4.4.1). In addition, we consider the quantitative estimates of exposure and risk provided 

by the HREA, as well as the uncertainties and limitations associated with these risk estimates 

(section 4.4.2). 
4-9 

4.4.1 Evidence-based Considerations 

In this section, we consider the available evidence from controlled human exposure and 

epidemiologic studies, including the uncertainties and limitations associated with that evidence, 

within the context of potential alternative standard levels. We consider both the exposure 

concentrations at which controlled human exposure studies provide evidence for health effects, 

and the ambient O3 concentrations present in locations where epidemiologic studies have 

reported health effect associations (see also section 3.1). 

Controlled human exposure studies and epidemiologic panel studies 

We consider the following question related to controlled human exposure studies and 

panel studies: 

 To what extent does the available evidence from controlled human exposure 

studies and panel studies provide support for consideration of potential 

alternative standard levels lower than 75 ppb? 

To inform our conclusions regarding this question, we consider the lowest O3 concentrations at 

which various effects have been evaluated and statistically significant effects reported. We also 

consider the potential for reported effects to be adverse, including in at-risk populations. 

As discussed in section 3.1.2.1, data from controlled human exposure studies show that 

group mean O3-induced lung function decrements in healthy adults exhibit a smooth doseresponse relationship without evidence of a threshold from 40 to 120 ppb O3 (US EPA, 2013, 

Figure 6-1). The lowest O3 exposure concentration for which statistically significant decrements 

have been reported is 60 ppb (Brown, 2008; Kim et al., 2011). The ISA concludes that mean 

FEV1 is clearly decreased by 6.6-hour exposures to O3 concentrations of 60 ppb and higher in 

young, healthy adults during moderate exertion (US EPA, 2013, p. 6-9). As discussed in section 

3.1.3, such a decrease in mean lung function meets the ATS criteria for an adverse response 

given that a downward shift in the distribution of FEV1 would result in diminished reserve 

function, and therefore would increase risk from further environmental insult. In addition, based 

on data from studies by Kim et al. (2011), Schelegle et al. (2009), Adams (2006), and Adams 

(1998), the ISA notes that following exposures to 60 ppb O3 10% of healthy adults experience 

FEV1 decrements > 10% (U.S. EPA, 2013, page 6-19).8 A 10% decrement in FEV1 is accepted 

 

8

As discussed in Chapter 3 of this PA (section 3.1.2.1), these estimates are consistent with the predictions 

of quantitative models developed by McDonnell et al. (2012) and Schelegle et al. (2012). The McDonnell 

model, as discussed in McDonnell et al. (2010), provides the basis for lung function risk estimates in the 

HREA (section 4.4.2.2, below). For the target of 60 ppb, Schelegle et al. (2009) reported an actual mean 

exposure concentration of 63 ppb. 
4-10 

by ATS as an abnormal response. Based on advice received from CASAC in this (Frey, 2014, p. 

3) and previous reviews, such decrements could be adverse in people with lung disease (section 

3.1.3). Moreover, as discussed in section 3.1.3 of this PA, repeated occurrences of moderate 

responses may be considered adverse since they could set the stage for more serious effects. 

One recent controlled human exposure study has reported O3-induced pulmonary 

inflammation (PMN increased in sputum from lower airways) following exposures of young, 

healthy adults to O3 concentrations of 60 ppb (Kim et al., 2011), the lowest concentration at 

which inflammatory responses have been evaluated in human studies (see discussion in section 

3.1.2.1). Induction of pulmonary inflammation is evidence that injury has occurred. The 

possibility of chronic effects due to repeated inflammatory events has been evaluated in animal 

studies. Repeated events of acute inflammation can have several potentially adverse outcomes 

including: induction of a chronic inflammatory state; altered pulmonary structure and function, 

leading to diseases such as asthma; altered lung host defense response to inhaled 

microorganisms, particularly in potentially at-risk populations such as the very young and old; 

and, altered lung response to other agents such as allergens or toxins (U.S. EPA, 2013, Section 

6.2.3). Thus, lung injury and the resulting inflammation, particularly if experienced repeatedly, 

provide a mechanism by which O3 may cause other more serious respiratory effects (e.g., asthma 

exacerbations) and possibly extrapulmonary effects. 

With respect to respiratory symptoms, a recent study by Schelegle et al. (2009) reported a 

statistically significant increase in respiratory symptoms in young, healthy adults following 6.6 

hour exposures to an average O3 concentration of 70 ppb.9

 This study also reported a statistically 

significant decrease in FEV1 following such exposures. As discussed in section 3.1.3, the 

occurrence of both lung function decrements and respiratory symptoms meets criteria established 

by the ATS defining an “adverse” respiratory response. Although some studies have reported 

that respiratory symptoms develop during exposures at 60 ppb, the increases in symptoms in 

these studies have not reached statistical significance by the end of the 6.6 hour exposures 

(Adams 2006; Schelegle et al., 2009).10 

Based on the results discussed above and in section 3.1.2.1, we conclude that controlled 

human exposure studies provide evidence of potentially adverse lung function decrements and 

airway inflammation in healthy adults following exposures to 60 ppb O3, and evidence of 

 

9

 For the target of 70 ppb, Schelegle et al. (2009) reported an actual mean exposure concentration of 72 ppb. 

10Adams (2006) reported an increase in respiratory symptoms in healthy adults during a 6.6 hour exposure protocol 

with an average O3 exposure concentration of 60 ppb. This increase was significantly different from initial 

respiratory symptoms, but not from the filtered air control day. For the target of 60 ppb, Schelegle et al. (2009) 

reported an actual mean exposure concentration of 63 ppb and did not observe a statistically significant increase in 

respiratory symptoms. 
4-11 

respiratory symptoms combined with lung function decrements (an “adverse” response based on 

ATS criteria) following exposures to 70 ppb. In reaching these conclusions, we recognize that 

most studies have not evaluated exposure concentrations below 60 ppb, and that 60 ppb does not 

necessarily reflect an exposure concentration below which effects no longer occur. Specifically, 

given the occurrence of airway inflammation in healthy adults following exposures to 60 ppb and 

higher, it may be reasonable to expect that inflammation would also occur following exposures 

to O3 concentrations somewhat below 60 ppb. Although some studies show that respiratory 

symptoms develop during exposures at 60 ppb, they have not reached statistical significance by 

the end of the 6.6 hour exposures (Adams 2006; Schelegle et al. 2009). Thus, respiratory 

symptoms combined with lung function decrements are likely to occur to some degree in healthy 

adults with 6.6-hour exposures to concentrations below 70 ppb, and are more likely to occur with 

8-hour exposures to 70 ppb and below. Further, we note that these controlled human exposure 

studies were conducted in healthy adults and that people with asthma, including asthmatic 

children, are likely to be more sensitive to O3-induced respiratory effects. Therefore, these 

exposure concentrations are more likely to cause adverse respiratory effects in children and 

adults with asthma, and more generally in people with respiratory disease. 

With regard to other O3-induced effects, we note that airway hyperresponsiveness and 

impaired lung host defense capabilities have been reported in healthy adults engaged in moderate 

exertion following exposures to O3 concentrations as low as 80 ppb, the lowest concentration 

evaluated for these effects.11 As discussed in section 3.1.2.1, these physiological effects have 

been linked to aggravation of asthma and increased susceptibility to respiratory infection, 

potentially leading to increased medication use, increased school and work absences, increased 

visits to doctors’ offices and emergency departments, and increased hospital admissions. These 

are all indicators of adverse O3-related morbidity effects, which are consistent with, and provide 

plausibility for, the adverse morbidity effects and mortality effects observed in epidemiologic 

studies. 

In further considering effects following exposures to O3 concentrations below 75 ppb, in 

section 3.1.4.1 we discuss panel studies highlighted in the ISA for the extent to which monitored 

ambient O3 concentrations reflect exposure concentrations in their study populations (U.S. EPA, 

2013, section 6.2.1.2). These panel studies used on-site monitoring to evaluate O3-attributable 

lung function decrements in people engaged in outdoor recreation, exercise, or work. Table 3-2 

includes O3 panel studies that report analyses of O3-attributable lung function decrements for O3

concentrations at or below 75 ppb, and that measure O3 concentrations with monitors located in 

the areas where study subjects were active (e.g., on site at summer camps or in locations where 

 

11There is no evidence that 80 ppb is a threshold for these effects (72 FR 37878, July 11, 2007). 
4-12 

exercise took place). Consistent with the results of controlled human exposure studies discussed 

above, these panel studies report associations with lung function decrements for subjects exposed 

to on-site monitored O3 concentrations below 75 ppb. Associations in panel studies have been 

reported for a wider range of populations than has been evaluated in controlled human exposure 

studies, including children. 

With regard to the question above, we conclude that the available controlled human 

exposure evidence and evidence from panel studies supports an upper end of the range of 

potential alternative standard levels for consideration no higher than 70 ppb. As just discussed, 

6.6-hour exposures of healthy adults to 70 ppb O3 result in lung function decrements and 

respiratory symptoms, a combination of effects that meet ATS criteria for an adverse response 

(as discussed in section 3.1.3).12 In addition, while 70 ppb is below the 80 ppb concentration 

shown in 6.6-hour exposure studies to cause potentially adverse respiratory effects such as 

airway hyperresponsiveness and impaired host-defense capabilities, these effects have not been 

evaluated at exposure concentrations below 80 ppb and there is no reason to believe that 80 ppb 

represents a threshold for such effects. As discussed in section 3.1.2.1 of this PA, the 

physiological effects reported in controlled human exposure studies down to 60 ppb O3 have 

been linked to aggravation of asthma and increased susceptibility to respiratory infection, 

potentially leading to increased medication use, increased school and work absences, increased 

visits to doctors’ offices and emergency departments, and increased hospital admissions. 

Based on the above considerations, we also conclude that the evidence from controlled 

human exposure studies and panel studies supports considering alternative O3 standard levels at 

least as low as 60 ppb. Potentially adverse lung function decrements and pulmonary 

inflammation have been demonstrated to occur in healthy adults at 60 ppb, with little evidence 

for potentially adverse effects following exposures to O3 concentrations below 60 ppb. Thus, 60 

ppb is a short-term exposure concentration that may be reasonably concluded to elicit adverse 

effects in at-risk groups. Pulmonary inflammation, particularly if experienced repeatedly, 

provides a mechanism by which O3 may cause other more serious respiratory morbidity effects 

(e.g., asthma exacerbations) and possibly extrapulmonary effects. 

Epidemiologic evidence 

We also consider what the information from epidemiologic studies indicates with regard 

to potential alternative standard levels appropriate for consideration. Based on the information in 

 

12 Based on the Schelegle et al. (2009) study, CASAC observed that, “adverse health effects in young healthy adults 

occur with exposures to 72 ppb of ozone for 6.6 hours” and that “It is the judgment of CASAC that if subjects had 

been exposed to ozone using the 8-hour averaging period used in the standard, adverse effects could have occurred 

at [a] lower concentration. Further, in our judgment, the level at which adverse effects might be observed would 

likely be lower for more sensitive subgroups, such as those with asthma” (Frey, 2014, p. 5). 
4-13 

section 3.1.4.2 of this PA (see Table 3-3), we first note that several epidemiologic studies have 

reported positive and statistically significant associations with hospital admissions, emergency 

department visits, and/or mortality in study areas where ambient O3 concentrations would have 

met the current standard (i.e., with its level of 75 ppb). This includes Canadian multicity studies 

in which the majority of study cities would have met the current standard over entire study 

periods (Cakmak et al., 2006; Dales et al., 2006; Katsouyanni et al., 2009; Stieb et al., 2009), and 

a U.S. single-city study conducted in a location likely to have met the current standard over the 

entire study period (Mar and Koenig, 2009). 

In further evaluating these studies, and building upon our conclusions based on controlled 

human exposures studies, as discussed above, we consider the following question related to the 

epidemiologic evidence: 

 To what extent have U.S. and Canadian epidemiologic studies reported associations 

with mortality or morbidity in locations likely to have met potential alternative O3

standards with levels from 70 to 60 ppb? 

Our focus in addressing this question is on what epidemiologic studies convey regarding the 

extent to which O3-associated health effects may be occurring (i.e., as indicated by associations) 

under air quality conditions allowed by potential alternative standards with levels of 70, 65, and 

60 ppb (Table 4-1).13 

 

13See ATA III, 283 F.3d at 370 (EPA justified in revising NAAQS when health effect associations are observed at 

levels allowed by the NAAQS). 
4-14 

Table 4-1 Numbers of epidemiologic study locations likely to have met potential 

alternative standards with levels of 70, 65, and 60 ppb 

Number of study cities meeting potential 

alternative standards during entire study 

period 

Study Result Cities 70 ppb 65 ppb 60 ppb 

Cakmak et al. 

(2006) 

Positive and statistically 

significant association with 

respiratory hospital 

admissions 

10 Canadian 

cities 7 6 2 

Dales et al. 

(2006) 

Positive and statistically 

significant association with 

respiratory hospital 

admissions 

11 Canadian 

cities 5 4 0 

Katsouyanni 

et al. (2009) 

Positive and statistically 

significant associations with 

respiratory hospital 

admissions 

12 Canadian 

cities 9 9 5 

Katsouyanni 

et al. (2009) 

Positive and statistically 

significant associations with 

total and cardiovascular 

mortality 

12 Canadian 

cities 7 5 1 

Mar and 

Koenig 

(2009) 

Positive and statistically 

significant associations with 

asthma emergency 

department visits 

Single city: 

Seattle 0 0 0 

Stieb et al. 

(2009) 

Positive and statistically 

significant association with 

respiratory emergency 

department visits 

7 Canadian 

cities 5 4 3 

As discussed in section 3.1.4.2, the single-city study by Mar and Koenig reported 

associations with respiratory emergency department visits in a location that would have met the 

current standard over the entire study period. In contrast, over at least part of the study period 

this area would have violated alternative O3 standards with levels of 70 ppb or below. Thus, 

while this study indicates that the current standard would allow the reported associations with 

respiratory emergency department visits, it does not provide information on the extent to which 

those health effect associations would be present if ambient O3 concentrations were reduced to 

meet a revised standard with a level at or below 70 ppb. 

With regard to the multicity studies included in Table 4-1, none were conducted in study 

locations that all would have met an O3 standard with a level at or below 70 ppb. However, for 

the studies by Cakmak et al. (2006), Katsouyanni et al. (2009), and Stieb et al. (2009), the 

majority of study locations would likely have met a standard with a level of either 70 or 65 ppb 

(Cakmak et al., 2006; Katsouyanni et al., 2009; Stieb et al., 2009). Thus the majority of the 

distributions of ambient O3 concentrations that provided the basis for positive and statistically 

significant associations with mortality or morbidity in these studies would likely be allowed 
4-15 

under alternative standards with levels of 70 or 65 ppb, though not 60 ppb. However, our 

interpretation of these results is complicated by uncertainties in the extent to which multicity 

effect estimates can be attributed to ambient O3 in the majority of locations, which would have 

met alternative standards, versus O3 in the smaller number of locations that would have violated 

those alternatives. 

As with our consideration of the current standard (section 3.1.4.2), we next consider the 

extent to which epidemiologic studies have characterized O3 health effect associations, including 

confidence in those associations, for various portions of distributions of ambient O3

concentrations. In considering such analyses within the context of potential alternative standards, 

we focus on the extent to which epidemiologic studies report health effect associations for air 

quality distributions restricted to ambient pollutant concentrations below one or more 

predetermined cut-points. As discussed in section 3.1.4.2, such “cut-point” analyses can provide 

information on the magnitude and statistical precision of effect estimates for defined 

distributions of ambient concentrations, which may in some cases include distributions that 

would be allowed by potential alternative standards. Specifically, we consider the following 

question: 

 To what extent do cut-point analyses from epidemiologic studies report health effect 

associations at ambient O3 concentrations that are likely to be allowed by potential 

alternative standards with levels from 70 to 60 ppb?

As with our consideration of the current standard in section 3.1.4.2 of this PA, we 

evaluate the cut-point analyses presented in the U.S. multicity study by Bell et al. (2006). These 

cut-point analyses can provide insights into the magnitude and statistical precision of health 

effect associations for different portions of the distribution of ambient concentrations, including 

insights into the ambient concentrations below which uncertainty in reported associations 

becomes notably greater. Our analysis of air quality data associated with the cut-points evaluated 

by Bell et al., and uncertainties associated with that analysis, is described elsewhere in this 

document (section 3.1.4.2). In this section, we consider what these cut-point analyses indicate 

with regard to the potential for health effect associations to extend to ambient O3 concentrations 

likely to be allowed by a revised O3 NAAQS with a level below 75 ppb. 

We particularly focus on the lowest cut-point for which the association between O3 and 

mortality was reported to be statistically significant (i.e., 30 ppb, as discussed in section 3.1.4.2). 

Based on the O3 air quality concentrations that met the criteria for inclusion in the 30 ppb cutpoint analysis, 84% of study areas had 3-year averages of annual 4th highest 8-hour daily max O3

concentrations at or below 70 ppb over the entire study period (Table 4-2). In addition, 64% of 

study areas had 3-year averages of annual 4th highest 8-hour daily max O3 concentrations at or 
4-16 

below 65 ppb (Table 4-2). In contrast, the majority of study areas had 4th highest concentrations 

above 60 ppb. 

Consistent with our interpretation of multicity effect estimates discussed above, these 

results suggest that the majority of the air quality distributions included in the 30 ppb O3 cut 

point would have been allowed by a standard with a level of 70 or 65 ppb. Thus the majority of 

the distributions of ambient O3 concentrations that provided the basis for a positive and 

statistically significant association with mortality would be allowed by alternative standards with 

levels of 70 or 65 ppb, but not 60 ppb. However, as discussed below our interpretation of these 

cut point analyses is complicated by important uncertainties. 

Table 4-2 Number of study cities with 3-year averages of 4th highest 8-hour daily max 

concentrations greater than 70, 65, or 60 ppb, for various cut-point analyses 

presented in Bell et al. (2006) 

Cut-point for 2-day moving average across monitors and cities (24-h avg)14

25 30 35 40 45 50 55 60 All 

Number (%) of 

Cities with 4th 

highest >70 (any 

3-yr period; 1987-

2000) 

0 (0%) 16 

(16%) 

55 

(56%) 

82 

(84%) 

89 

(91%) 

92 

(94%) 

94 

(96%) 

95 

(97%) 

95 

(97%) 

Number (%) of 

Cities with 4th 

highest >65 (any 

3-yr period; 1987-

2000) 

3 (3%) 35 

(36%) 

77 

(79%) 

89 

(91%) 

94 

(96%) 

95 

(97%) 

95 

(97%) 

95 

(97%) 

95 

(97%) 

Number (%) of 

Cities with 4th 

highest >60 (any 

3-yr period; 1987-

2000) 

16 

(16%) 

61 

(62%) 

86 

(88%) 

94 

(96%) 

95 

(97%) 

96 

(8%) 

96 

(8%) 

96 

(8%) 

96 

(8%) 

In further considering the implications of Tables 4-1 and 4-2 for potential alternative 

standard levels, we also note the important uncertainties described in section 3.1.4 of this 

document. General uncertainties include the geographic heterogeneity in effect estimates, which 

could obscure presence of potential thresholds in multicity studies; uncertainty in the extent to 

which multicity effect estimates can be attributed to ambient O3 in the majority of locations, 

which would have met alternative standards with levels of 70 or 65 ppb, versus O3 in the smaller 

number of locations that would have violated those alternatives; and uncertainty in the extent to 

which the relatively low ambient O3 concentrations present in some study areas caused or 

 

14Cut point analyses presented in the study by Bell et al. (2006) are described in more detail in sections 3.1.2.3 and 

3.1.4.2 of this document. 
4-17 

contributed to reported effects. Additional uncertainties specific to our analysis of the cut points 

presented by Bell et al. (2006) include the appropriateness of identifying 4th highest 

concentrations from air quality subsets, rather than the entire air quality distributions that existed 

in study locations, and uncertainty associated with the air quality data used to re-create the cutpoint analyses from the published study. With regard to this second uncertainty, as described in 

more detail in section 3.1.4.2 of this document, our re-creation of the cut points was based on air 

quality data available in AQS, combined with the published descriptions of cut point criteria and 

study area definitions. In doing so, we did not recreate the trimmed means used by Bell. 

Therefore, an important uncertainty in this approach is the extent to which we were able to 

appropriately re-create the cut-point analyses in the published study. 

Overall, our analyses of air quality in U.S. and Canadian epidemiologic study locations 

indicate that (1) single-city studies have not been conducted in locations that would have met 

alternative O3 standards with levels of 70 ppb or below and that (2) multicity epidemiologic 

studies report positive and statistically significant associations with mortality and morbidity 

based largely on distributions of ambient O3 that would have been allowed by alternative 

standards with levels of 70 or 65 ppb, but not 60 ppb. While important uncertainties, mentioned 

above, complicate our interpretation of the multicity studies, at a minimum these results suggest 

that an alternative standard level of 60 ppb would not allow the distributions of ambient O3

concentrations present in the majority of study locations that provided the basis for statistically 

significant health effect associations. While the potential implications for alternative standard 

levels of 70 and 65 ppb are less clear, given the important uncertainties in these analyses, the 

results suggest that positive and statistically significant associations with mortality or morbidity 

in some studies were largely influenced by air quality distributions that would be allowed under 

alternative standards with such levels. 

We next consider the extent to which epidemiologic studies employing longer-term 

ambient O3 concentration metrics can inform our consideration of potential alternative standard 

levels. In doing so, we consider the following question: 

 To what extent does the available evidence indicate that an O3 standard with a 

level from 70 to 60 ppb, combined with the current 8-hour averaging time and 4th

high form, could provide protection from long-term exposures to ambient O3

concentrations for which there is evidence of health effects? 

We first note that, as discussed in section 3.1.4.3 of this PA, virtually all of the study 

cities that provided the basis for the positive and statistically significant association between 

long-term O3 and respiratory mortality (Jerrett et al., 2009) would have violated the current 

standard, and therefore potential alternative standards with lower levels. Thus, as with our 
4-18 

consideration of the current standard in section 3.1.4.3, while the study by Jerrett et al. (2009) 

contributes to our understanding of health effects associated with ambient O3 (summarized in 

section 3.1.2), it is less informative regarding the extent to which those health effects may be 

occurring under air quality conditions that would meet potential alternative standards. 

To further evaluate this issue, we use the adjusted air quality in urban case study areas, as 

described in the HREA, to consider the extent to which just meeting alternative O3 standards 

with levels of 70, 65, and 60 ppb could maintain long-term O3 concentrations below those in the 

cities that provided the basis for the positive and statistically significant association with 

respiratory mortality reported by Jerrett et al. (2009).15 Upon adjustment of air quality in U.S. 

urban case study areas to meet the current and potential alternative 8-hour standards, seasonal 

average 1-hour daily max concentrations were calculated and compared to the concentrations in 

study cities. 

As discussed in section 3.1.4.3, Jerrett et al. (2009) reported that when seasonal averages 

of 1-hour daily max O3 concentrations16 ranged from 33 to 104 ppb, there was no statistical 

deviation from a linear concentration-response relationship between O3 and respiratory mortality 

across 96 U.S. cities (U.S. EPA, 2013, section 7.7). However, as discussed in section 3.1.4.3, the 

study suggests notably decreased confidence in the reported linear concentration-response 

function for “long-term” O3 concentrations in the first quartile (i.e., at or below about 53 ppb), 

given the widening in confidence intervals for lower concentrations (based on visual inspection 

of Figure 3-6 in section 3.1.4.3); the fact that most study cities contributing to the linear function 

had O3 concentrations in the highest three quartiles, accounting for approximately 72% of the 

respiratory deaths in the cohort (based on Table 2 in the published study); and the limited 

evidence presented in the published study for a threshold at or near 56 ppb.17 

Given the above, we note the extent to which long-term O3 concentrations (i.e., seasonal 

average of 1-hour daily max) in urban case study areas are estimated to be at or below 53 ppb 

following air quality adjustment to meet potential alternative standards with levels of 70, 65, and 

60 ppb. To the extent air quality adjustment to just meet potential alternative short-term 

standards results in long-term concentrations near or below 53 ppb, we have greater confidence 

 

15Air quality in U.S. urban case study areas was adjusted to just meet the current 8-hour standard at 75 ppb, as well 

as potential potential alternative 8-hour standards at 70 ppb, 65 ppb, and 60 ppb, as described in the HREA (chapter 

4). After a given adjustment, seasonal average 1-hour daily max concentrations were calculated. 

16Jerrett et al. (2009) evaluated the April to September averages of 1-hour daily max O3 concentrations across 96 

U.S. metropolitan areas from 1977- 2000. In urban areas with multiple monitors, April to September 1-hour daily 

max concentrations from each individual monitor were averaged. This step was repeated for each year in the study 

period. Finally, each yearly averaged O3 concentrations was then averaged again to yield the single averaged 1-hour 

daily max O3 concentration depicted on the x-axis of Figure 3-6 below. 

17The issue of potential thresholds based on the Jerrett study is discussed in more detail in section 3.2.3.2 of this PA. 
4-19 

in the degree to which those short-term standards could protect against the health effects 

associated with longer term O3 exposures. Though there is uncertainty associated with these 

comparisons (e.g., due to uncertainty in the potential for a threshold to exist; uncertainty in the 

identification of such a threshold, should one exist; uncertainty in the long-term concentration 

below which confidence intervals widen notably, based on visual inspection of concentrationresponse function in the published study; and the limited number of urban case study areas for 

which adjusted air quality is available), this analysis can provide insight into the extent to which 

various alternative short-term standards would be expected to maintain long-term O3

concentrations below those where we have the most confidence in the reported concentrationresponse relationship with respiratory mortality. 

Table 4-3 indicates that when considering recent (i.e., unadjusted) air quality, 2 of 12 

urban case study areas had seasonal average 1-hour daily max O3 concentrations at or below 53 

ppb in all of the years examined. When air quality was adjusted to just meet the current 8-hour 

standard (75 ppb in Table 4-3), 6 of 12 urban case study areas had seasonal average 1-hour daily 

max O3 concentrations at or below 53 ppb in all of the years examined. When air quality is 

further adjusted to just meet potential alternative standards with lower levels, seasonal averages 

of 1-hour daily max O3 concentrations are estimated to be at or below 53 ppb in 9 of 12 urban 

case study areas (70 ppb level), 10 of 12 urban case study areas (65 ppb level), and 11 of 11 

urban case study areas (60 ppb level).18 Though as noted above there are important uncertainties 

associated with interpreting these comparisons, they suggest that in many locations across the 

U.S. a standard with an 8-hour averaging time, when combined with the current 4th high form 

and an appropriate standard level, would be expected to maintain seasonal averages of 1-hour 

daily max O3 concentrations below those where analyses indicate the most confidence in the 

concentration-response relationship with respiratory mortality reported by Jerrett et al. (2009). 

 

18As described in the HREA, a standard level of 60 ppb was not evaluated in New York City (U.S. EPA, 2014, 

chapter 4). 
4-20 

Table 4-3 Seasonal averages of 1-hour daily max O3 concentrations in U.S. urban case 

study areas for recent air quality and air quality adjusted to just meet the 

current and potential alternative standards. 

Air Quality

Adjusted to:

2006                 

(Adj Yrs 2006‐2008)

2007                

(Adj Yrs 2006‐2008)

2008                

(Adj Yrs 2008‐2010)

2009                  

(Adj Yrs 2008‐2010)

2010                 

(Adj Yrs 2008‐2010)

Recent 65 63 57 50 56

75 53 52 53 47 52

70 50 49 49 44 49

65 47 46 46 42 46

60 45 44 44 40 44

Recent 60 59 57 52 60

75 54 54 53 49 55

70 52 51 51 48 53

65 49 49 48 46 50

60 46 46 46 44 48

Recent 49 50 46 45 49

75 48 49 49 45 48

70 46 47 48 44 48

65 44 45 46 43 46

60 43 43 44 41 44

Recent 51 52 53 49 54

75 49 50 51 47 51

70 47 48 48 45 48

65 45 45 45 43 45

60 41 41 41 40 42

Recent 63 63 63 58 60

75 62 61 63 58 60

70 60 59 62 58 58

65 58 58 59 56 55

60 53 53 53 51 50

Recent 50 54 51 48 52

75 50 52 NA NA NA

70 48 50 51 49 52

65 47 49 49 47 50

60 45 46 46 45 47

Recent 53 48 47 47 46

75 48 46 47 48 46

70 47 45 46 47 46

65 46 44 45 46 45

60 45 43 43 44 44

Recent 65 61 64 62 57

75 58 59 60 60 58

70 55 56 57 58 56

65 52 53 54 54 53

60 50 51 52 52 50

Recent 53 54 55 48 55

75 47 47 51 47 51

70 44 45 48 45 48

65 36 36 39 38 39

60 NA NA NA NA NA

Recent 56 59 57 51 58

75 51 52 54 49 54

70 49 50 51 47 52

65 47 48 49 45 49

60 45 46 47 43 47

Recent 66 59 65 61 55

75 55 50 54 51 48

70 52 48 51 49 46

65 50 46 49 47 44

60 47 44 46 44 42

Recent 58 58 52 51 55

75 53 53 51 50 54

70 50 51 50 48 52

65 47 48 48 46 49

60 44 45 45 43 46

Houston

Philadelphia

Sacramento

Saint Louis

Atlanta

Baltimore

Boston

Cleveland

Denver

Detroit

Los Angeles

New York City
4-21 

Based on the above analyses, we conclude that the available epidemiologic evidence is 

consistent with the available evidence from controlled human exposure studies in providing 

support for consideration of an O3 standard level in the range of 70 to 60 ppb. Compared to the 

current standard, a standard level from within this range would expected to be more effective at 

maintaining short-term and long-term ambient O3 concentrations below those present in studies 

reporting O3-associated mortality and/or morbidity. 

In reaching overall staff conclusions about an appropriate range of standard levels for 

consideration, we further evaluate the results of the exposure and risk assessments that are based 

on modeling changes in the entire distribution of ambient O3 concentrations to simulate just 

meeting potential alternative standards. These results are discussed below in section 4.4.2. 

4.4.2 Air Quality-, Exposure-, and Risk-Based Considerations 

Beyond considering the available evidence, we also consider the extent to which specific 

potential alternative standard levels, in conjunction with the current averaging time and form (3-

year average of annual 4th highest 8-hour daily max), could reduce estimated O3 exposures and 

health risks. In the first draft PA (U.S. EPA, 2012b), we concluded that the available evidence 

supports conducting further exposure and risk analyses of potential alternative O3 standard levels 

in the range of 70 down to 60 ppb. Based on these conclusions, the HREA evaluates exposures 

and risks estimated to be associated with potential alternative standard levels from the upper (70 

ppb), middle (65 ppb), and lower (60 ppb) portions of this range. In considering these analyses in 

this PA, we consider the following question: 

 To what extent does the available exposure and risk information provide support 

for considering potential alternative standard levels from 70 to 60 ppb, when 

combined with the current 8-hour averaging time and 4th high form? 

In considering exposure and risk analyses, we emphasize the nature and magnitude of the O3

exposures and health risks estimated to remain upon just meeting each alternative standard level, 

and the changes in exposures and risks estimated for each alternative level when compared to the 

current standard. Section 4.4.2.1 below discusses our exposure-based considerations. Sections 

4.4.2.2 and 4.4.2.3 discuss our consideration of estimates of lung function risks and estimates of 

epidemiology-based mortality/morbidity risks, respectively. 

4.4.2.1 Exposure-Based Considerations 

As discussed in more detail in section 3.2.2 of this PA, the exposure assessment 

presented in the HREA (U.S. EPA, 2014, Chapter 5) provides estimates of the number and 

percent of people exposed to O3 concentrations at or above benchmark concentrations of 60, 70, 

and 80 ppb, while at moderate or greater exertion. Estimates of such “exposures of concern” 
4-22 

provide perspective on the potential public health impacts of O3-related effects, including for 

effects that cannot currently be evaluated in a quantitative risk assessment. The approach taken 

in the HREA to estimating exposures of concern, and the key uncertainties associated with 

exposure estimates, are summarized in section 3.2.2 for air quality adjusted to just meet the 

current standard and are discussed in more detail in chapter 5 of the HREA (U.S. EPA, 2014). As 

discussed in section 3.2.2, when evaluating potential alternative standard levels we focus on 

modeled exposures for school-age children (ages 5-18), noting that percentages of asthmatic 

school-age children estimated to experience exposures of concern are virtually indistinguishable 

from those for all children, and that patterns of exposure in children represent a broader range of 

at-risk populations, which includes adult asthmatics and older adults. In this review, CASAC 

advised EPA to focus on the 60 ppb benchmark as being relevant for considering adverse effects 

on people with asthma (Frey, 2014, p. 6). 

In this section, we consider the following question: 

 To what extent are potential alternative standards with revised levels estimated to 

reduce the occurrence of O3 exposures of concern, compared to the current 

standard, and what are the nature and magnitude of the exposures remaining for 

each alternative standard level evaluated? 

Key results related to this question are summarized below (Figures 4-1 to 4-4). Figures 4-1 

(estimates averaged over years) and 4-2 (estimates from worst-case years) present estimates of 

one or more exposures of concern, and Figures 4-3 (estimates averaged over years) and 4-4 

(estimates from worst-case years) present estimates of two or more exposures of concern. 
4-23 

Figure 4-1. Percent of children estimated to experience one or more exposures of concern at or above 60, 

quality adjusted to just meet the current and potential alternative standards (averaged over 2
4-24 

Figure 4-2. Percent of children estimated to experience one or more exposures of concern at or above 60, 

quality adjusted to just meet the current and potential alternative standards (worst-case year 201019) 

19“Worst-case” year refers to the year in each urban case study area with the largest percentage of children estimated to experience expo
4-25 

Figure 4-3. Percent of children estimated to experience two or more exposures of concern at or above 60, 

quality adjusted to just meet the current and potential alternative standards (averaged over 2
4-26 

Figure 4-4. Percent of children estimated to experience two or more exposures of concern at or above 60, 

quality adjusted to just meet the current and potential alternative standards (worst-case year 
4-27 

As illustrated above in Figures 4-1 to 4-4, adjusting air quality to just meet progressively 

lower potential alternative standard levels reduces estimated exposures of concern consistently 

across urban case study areas. These results reflect the consistent reductions in the highest 

ambient O3 concentrations upon air quality adjustment, as summarized in section 3.2.1 and as 

discussed in more detail in the HREA (U.S. EPA, 2014, chapter 4). Based on Figures 4-1 to 4-4 

and the associated details described in the HREA (U.S. EPA 2014, chapter 5), we take note of 

the following with regard to exposures of concern for specific potential alternative standard 

levels: 

1. For an O3 standard level of 70 ppb: 

a. On average over the years 2006 to 2010, a standard with a level of 70 ppb is 

estimated to allow approximately 3 to 10% of children in urban case study areas to 

experience one or more exposures of concern at or above 60 ppb (approximately 30 to 

70% reduction, relative to current standard). Summing across urban case study areas, 

these percentages correspond to over 1 million children experiencing over 1.5 million 

exposures of concern at or above 60 ppb during a single O3 season. Of these children, 

over 100,000 are asthmatics. 

b. On average over the years 2006 to 2010, a standard with a level of 70 ppb is 

estimated to allow approximately 0.5 to 3.5% of children in urban case study areas to 

experience two or more exposures of concern at or above 60 ppb (approximately 50 

to 85% reduction, relative to current standard). 

c. In the worst-case years (i.e., those with the largest exposure estimates), a standard 

with a level of 70 ppb is estimated to allow approximately 5 to 19% of children in 

urban case study areas to experience one or more exposures of concern at or above 60 

ppb, and approximately 2 to 9% to experience two or more. 

d. On average over the years 2006 to 2010, a standard with a level of 70 ppb is 

estimated to allow approximately 1% or less of children to experience one or more 

exposures of concern at or above 70 ppb (approximately 55 to 90% reduction, relative 

to current standard), and far less than 1% to experience two or more such exposures 

(approximately 65 to 100% reduction, relative to current standard). 

e. In the worst-case years, approximately 3% or less of children are estimated to 

experience one or more exposures of concern at or above 70 ppb, and less than 1% 

are estimated to experience two or more such exposures. 

f. A standard with a level of 70 ppb is estimated to allow less than 1% of children to 

experience one or more exposures of concern at or above 80 ppb, even in the worstcase years. No children are estimated to experience two or more such exposures. 

2. For an O3 standard level of 65 ppb: 
4-28 

a. On average over the years 2006 to 2010, a standard with a level of 65 ppb is 

estimated to allow approximately 4% or less of children in urban case study areas to 

experience one or more exposures of concern at or above 60 ppb (approximately 70 to 

100% reduction, relative to current standard). Summing across urban case study 

areas, these percentages correspond to almost 400,000 children experiencing almost 

500,000 exposures of concern at or above 60 ppb during a single O3 season. Of these 

children, about 40,000 are asthmatics. 

b. On average over the years 2006 to 2010, a standard with a level of 65 ppb is 

estimated to allow less than 1% of children to experience two or more exposures of 

concern at or above 60 ppb (approximately 85 to 100% reduction, relative to current 

standard). 

c. In the worst-case years, a standard with a level of 65 ppb is estimated to allow 

approximately 10% or less of children to experience one or more exposures of 

concern at or above 60 ppb, and approximately 3% or less to experience two or more 

such exposures. 

d. On average over the years 2006 to 2010, a standard with a level of 65 ppb is 

estimated to allow approximately 1% or less of children to experience one or more 

exposures of concern at or above 70 ppb (approximately 90 to 100% reduction, 

relative to current standard), and almost no children to experience two or more such 

exposures. Even in the worst-case years, a level of 65 ppb is estimated to allow less 

than 1% of children to experience exposures of concern at or above 70 ppb. 

e. A standard with a level of 65 ppb is estimated to allow virtually no children to 

experience exposures of concern at or above 80 ppb, even in the worst-case years. 

3. For an O3 standard level of 60 ppb: 

a. On average over the years 2006 to 2010, a standard with a level of 60 ppb is 

estimated to allow approximately 1% or less of children to experience one or more 

exposures of concern at or above 60 ppb (approximately 90 to 100% reduction, 

relative to current standard), and virtually no children to experience multiple such 

exposures. 

b. In the worst-case years, a standard with a level of 60 ppb is estimated to allow 

approximately 2% or less of children to experience one or more exposures of concern 

at or above 60 ppb, and almost no children to experience multiple such exposures. 

c. On average over the years 2006 to 2010, a standard with a level of 60 ppb is 

estimated to almost eliminate exposures of concern at or above 70 ppb or 80 ppb. 

Even in years with the highest exposure estimates, virtually no children are estimated 

to experience such exposures. 

In further considering these exposure estimates, we take note of the associated 

uncertainties, as discussed in more detail in section 3.2.2 of this PA (and in Chapter 5 of the 
4-29 

HREA, U.S. EPA, 2014). These include (1) individual variability in responsiveness to O3

exposures such that only a subset of individuals who experience exposures at (or above) a 

benchmark concentration would experience health effects; (2) potential to underestimate 

exposures in most highly exposed populations; and (3) potential to overestimate exposures in 

populations who alter behavior in response to high O3 days (i.e., spend less time being active 

outdoors). The implications of estimated exposures of concern for potential alternative standard 

levels are discussed below in section 4.6. 

4.4.2.2 Risk-Based Considerations: Lung Function 

As discussed above in more detail in section 3.2.3.1 of this PA, the assessment of lung 

function risks presented in the HREA (U.S. EPA, 2014, Chapter 6) provides estimates of the 

number and percent of people experiencing O3-induced lung function decrements greater than or 

equal to 10, 15, and 20%. In the current and past reviews, CASAC has advised EPA to focus on 

decrements of 10% or greater when considering people with pre-existing lung disease (Frey, 

2014; Samet, 2011). 

Lung function risk estimates are based on an updated dose-threshold model that estimates 

FEV1 responses for healthy adults following short-term exposures to O3 (McDonnell, Stewart, 

and Smith, 2010), reflecting methodological improvements since the last review (U.S. EPA, 

2014, section 6.2.4). The approach taken in the HREA to estimating O3-induced lung function 

decrements, and the key uncertainties associated with these estimates, are summarized in section 

3.2.3.1 for air quality adjusted to just meet the current standard and are discussed in more detail 

in chapter 6 of the HREA (U.S. EPA, 2014). 

As discussed in section 3.2.3.1, in evaluating potential alternative standard levels we 

focus on modeled exposures for school-age children, with an emphasis on asthmatic children. As 

with exposures of concern, the percentages of all school age children and asthmatic school age 

children estimated to experience particular O3-induced lung function decrements are virtually 

indistinguishable. 

In this section, we consider the following question: 

 To what extent are potential alternative standards with revised levels estimated to 

decrease the occurrence of O3-induced lung function decrements, compared to 

the current standard, and what are the nature and magnitude of the decrements 

remaining for each alternative standard level evaluated? 

Key results related to this question are summarized below (Figures 4-5 to 4-8). Figures 4-5 

(estimates averaged over years) and 4-6 (estimates from worst-case years) present estimates of 

one or more O3-induced lung function decrements, and Figures 4-7 (estimates averaged over 

years) and 4-8 (estimates from worst-case years) present estimates of two or more decrements. 
4-30 

Figure 4-5. Percent of children estimated to experience one or more O3-induced lung function decrement

or 20% for air quality adjusted to just meet the current and potential alternative standards (ato 2010) 
4-31 

Figure 4-6. Percent of children estimated to experience one or more O3-induced lung function decrement

or 20% for air quality adjusted to just meet the current and potential alternative standards (w2006 to 2010) 
4-32 

Figure 4-7. Percent of children estimated to experience two or more O3-induced lung function decrement

15, or 20% for air quality adjusted to just meet the current and potential alternative standard2006 to 2010) 
4-33 

Figure 4-8. Percent of children estimated to experience two or more O3-induced lung function decrement

15, or 20% for air quality adjusted to just meet the current and potential alternative standardfrom 2006 to 2010) 
4-34 

As illustrated above in Figures 4-5 to 4-8, adjusting air quality to just meet progressively 

lower potential alternative standard levels consistently reduces the percent of children estimated 

to experience potentially adverse lung function decrements. These results reflect the consistent 

reductions in the highest ambient O3 concentrations upon air quality adjustment (section 3.2.1; 

U.S. EPA, 2014, chapter 4).20 Based on Figures 4-5 to 4-8 and the associated details described in 

the HREA (U.S. EPA 2014, chapter 6), we take note of the following with regard to specific 

potential alternative standard levels: 

1. For an O3 standard level of 70 ppb: 

a. On average over the years 2006 to 2010, a standard with a level of 70 ppb is 

estimated to allow approximately 11 to 17% of children in urban case study areas, 

including asthmatic children, to experience one or more O3-induced lung function 

decrements ≥ 10% (approximately 6 to 27% reduction, relative to current 

standard) per season. Summing across case study areas, these percentages 

correspond to approximately 260,000 asthmatic children experiencing 

approximately 1 million total occurrences of O3-induced lung function 

decrements greater than or equal to 10%. 

b. On average over the years 2006 to 2010, a standard with a level of 70 ppb is 

estimated to allow approximately 6 to 11% of children, including asthmatic 

children, to experience two or more O3-induced lung function decrements ≥ 10% 

(approximately 8 to 30% reduction, relative to current standard). 

c. In the worst-case years, a standard with a level of 70 ppb is estimated to allow 

approximately 14 to 20% of children, including asthmatic children, to experience 

one or more O3-induced lung function decrements ≥10%, and approximately 7 to 

13% to experience two or more such decrements. 

d. On average over the years 2006 to 2010, a standard with a level of 70 ppb is 

estimated to allow approximately 2 to 4% of children, including asthmatic 

children, to experience one or more O3-induced lung function decrements ≥ 15%, 

and approximately 1 to 2.5% of children to experience two or more such O3-

induced decrements. In the worst-case years, approximately 3 to 5% of children 

are estimated to experience one or more O3-induced lung function decrements 

≥15%, and approximately 1 to 3% are estimated to experience two or more such 

decrements. 

e. A standard with a level of 70 ppb is estimated to allow 2% or fewer children to 

experience any O3-induced lung function decrements ≥ 20%, even in the worstcase years. Approximately 1% or fewer children are estimated to experience two 

 

20As discussed in section 3.2.3.1, the impact of the dose threshold in the lung function risk model is that O3-induced 

FEV1 decrements result primarily from exposures on days with average ambient O3 concentrations above about 40 

ppb (US EPA, 2014, section 6.3.1, Figure 6-9). 
4-35 

or more O3-induced lung function decrements ≥ 20%, even in the worst-case 

years. 

2. For an O3 standard level of 65 ppb: 

a. On average over the years 2006 to 2010, a standard with a level of 65 ppb is 

estimated to allow approximately 3 to 15% of children, including asthmatic 

children, to experience one or more O3-induced lung function decrements ≥ 10% 

(approximately 20 to 77% reduction, relative to current standard). Summing 

across urban case study areas, these percentages correspond to approximately 

190,000 asthmatic children experiencing almost 750,000 total occurrences of O3-

induced lung function decrements ≥ 10%. 

b. On average over the years 2006 to 2010, a standard with a level of 65 ppb is 

estimated to allow approximately 1 to 9% of children, including asthmatic 

children, to experience two or more O3-induced lung function decrements ≥ 10% 

(approximately 20 to 80% reduction, relative to current standard). 

c. In the worst-case years, a standard with a level of 65 ppb is estimated to allow 

approximately 4 to 18% of children to experience one or more O3-induced lung 

function decrements ≥ 10%, and approximately 2 to 11% to experience two or 

more such decrements. 

d. On average over the years 2006 to 2010, a standard with a level of 65 ppb is 

estimated to allow approximately 3% or less of children to experience one or 

more O3-induced lung function decrements ≥ 15%, and approximately 2% or less 

of children to experience two or more such O3-induced decrements. In the worstcase years, approximately 4% or less of children are estimated to experience one 

or more O3-induced lung function decrements ≥ 15%, and up to approximately 

2% are estimated to experience two or more such decrements. 

e. A standard with a level of 65 ppb is estimated to allow less than 1.5% of children 

to experience any O3-induced lung function decrements ≥ 20%, even in the worstcase years. A standard with a level of 65 ppb is estimated to allow less than 1% of 

children to experience two or more O3-induced lung function decrements ≥ 20%, 

even in the worst-case years. 

3. For an O3 standard level of 60 ppb: 

a. On average over the years 2006 to 2010, a standard with a level of 60 ppb is 

estimated to allow approximately 5 to 11% of children, including asthmatic 

children, to experience one or more O3-induced lung function decrements ≥ 10% 

(approximately 35 to 77% reduction, relative to current standard). Summing 

across urban case study areas, these percentages correspond to approximately 

140,000 asthmatic children experiencing approximately 500,000 total occurrences 

of O3-induced lung function decrements ≥ 10%. 
4-36 

b. On average over the years 2006 to 2010, a standard with a level of 60 ppb is 

estimated to allow approximately 2 to 6% of children to experience two or more 

O3-induced lung function decrements ≥ 10% (approximately 40 to 70% reduction, 

relative to current standard). 

c. In the worst-case years, a standard with a level of 60 ppb is estimated to allow 

approximately 5 to 13% of children to experience one or more O3-induced lung 

function decrements ≥ 10%, and approximately 2 to 7% to experience two or 

more such decrements. 

d. A standard with a level of 60 ppb is estimated to allow less than about 3% of 

children to experience any O3-induced lung function decrements ≥ 15% and less 

than 1% to experience decrements greater than 20%, even in the worst-case years. 

A standard with a level of 60 ppb is estimated to allow less than 1.5% of children 

to experience two or more O3-induced lung function decrements ≥ 15% and less 

than 0.5% to experience two or more decrements ≥ 20%, even in the worst-case 

years. 

In further considering these exposure estimates, we take note of the associated 

uncertainties, as discussed in more detail in section 3.2.2 of this PA. In addition to the 

uncertainties in exposure estimates noted above, these include the relative lack of exposureresponse information for key at-risk populations (i.e., children and asthmatics), since most 

controlled human exposures studies are conducted in healthy adults. Section 4.6 (below) 

discusses the implications of estimates of the occurrence of O3-induced lung function decrements 

for potential alternative standard levels. 

4.4.2.3 Risk-Based Considerations: Epidemiology-Based Mortality and Morbidity 

The epidemiology-based risk assessments presented in the HREA (U.S. EPA, 2014, 

chapter 7) provide estimates of total mortality, respiratory hospital admissions and emergency 

department visits, and asthma exacerbations associated with short-term O3 concentrations. The 

HREA also presents estimates of respiratory mortality associated with long-term21

concentrations. In evaluating these risk estimates, we consider the following question: 

 

21Estimates of respiratory mortality associated with long-term O3 concentrations are based on the study by Jerrett et 

al. (2009). Consistent with the O3 metric used in the study, risk estimates are based on seasonal averages of 1-hour 

daily max O3 concentrations. 
4-37 

 To what extent are potential alternative standards with revised levels estimated to 

decrease O3 health risks, compared to the current standard, and what are the 

nature and magnitude of the health risks remaining for each alternative standard 

level evaluated? 

As discussed in more detail in section 3.2.3.2 of this PA, in considering this question we 

are mindful that the model-based approach used to adjust air quality in the HREA has important 

implications for risk estimates developed by applying concentration-response relationships from 

epidemiologic studies (section 3.2.1). In particular, given the use of linear concentrationresponse relationships, risk estimates are equally influenced by decreasing high O3

concentrations and increasing low O3 concentrations following air quality adjustment, when the 

increases and decreases are of equal magnitude. This and other uncertainties associated with risk 

estimates are discussed in section 3.2.3.2. 

Key results from the HREA (U.S. EPA, 2014, chapter 7) are summarized below for 

estimates of total mortality associated with short-term O3 concentrations (Figures 4-9 and 4-10) 

and respiratory hospital admissions associated with short-term O3 concentrations (Figure 4-11). 

The other morbidity effects evaluated in the HREA (i.e., respiratory emergency department visits 

and asthma symptoms associated with short-term concentrations) exhibit patterns across standard 

levels that are similar to those reported for total mortality and respiratory hospital admissions 

(U.S. EPA, 2014, chapter 7). 
4-38 

Figure 4-9. Estimates of Total Mortality Associated with Short-Term O3 Concentrations 

in Urban Case Study Areas (Air Quality Adjusted to Current and Potential 

alternative standard levels) – Total Risk 

The risk estimates presented in Figure 4-9 above are based on applying linear 

concentration-response relationships to the full distributions of daily 8-hour “area-wide” O3

concentrations. However, as in section 3.2.3.2 we note the ISA conclusion that there is less 

certainty in the shape of concentration-response functions for area-wide O3 concentrations at the 

lower ends of warm season distributions (i.e., below about 20 to 40 ppb depending on the O3

2007 Simulation year

2009 Simulation year

0

2

4

6

8

10

12

14

16

18

20

75ppb 70ppb 65ppb 60ppb

Total ozone‐related mortality per 100,000 residents

Trend in ozone‐related mortality across standard

levels (deaths per 100,000)

Atlanta, GA

Baltimore, MD

Boston, MA

Cleveland, OH

Denver, CO

Detroit, MI

Houston, TX

Los Angeles, CA

New York, NY

Philadelphia, PA

Sacramento, CA

St. Louis, MO

0

2

4

6

8

10

12

14

16

18

20

75ppb 70ppb 65ppb 60ppb

Total ozone‐related mortality per 100,000 residents

Trend in ozone‐related mortality across standard levels

(deaths per 100,000)

Atlanta, GA

Baltimore, MD

Boston, MA

Cleveland, OH

Denver, CO

Detroit, MI

Houston, TX

Los Angeles, CA

New York, NY

Philadelphia, PA

Sacramento, CA

St. Louis, MO
4-39 

metric, health endpoint, and study population) (U.S. EPA, 2013, section 2.5.4.4). We also 

recognize that for the range of health endpoints evaluated, controlled human exposure and 

animal toxicological studies provide greater certainty in the increased incidence, magnitude, and 

severity of effects at higher exposure concentrations (discussed in sections 3.1.2.2 and 3.1.4.2 of 

this document).22 Thus, in addition to considering estimates of total O3-associated risks, we also 

consider the extent to which risks are associated with days with higher, versus lower, area-wide 

O3 concentrations. 

Figure 4-10 presents estimates of O3-associated deaths, summed across urban case study 

areas, for days with area-wide concentrations at or above 20, 40, and 60 ppb. As discussed in 

more detail in section 3.2.1 of this document, daytime O3 concentrations in the upper portions of 

the distributions of area-wide concentrations tend to decrease upon adjustment to meet lower 

potential alternative standard levels, while concentrations in the lower portions of these 

distributions tend to increase. As a result, lower standard levels are estimated to be more 

effective at reducing deaths associated with the upper portions of these distributions of ambient 

O3 concentrations than deaths associated with the full distributions.23 

 

22As discussed in section 3.1.4.2, as ambient concentrations increase the potential for exposures to higher O3

concentrations also increases. Thus with increasing ambient concentrations, controlled human exposure and animal 

toxicological studies provide greater certainty in the increased incidence, magnitude, and severity of O3-attributable 

effects. 

23The relatively small proportion of O3-associated deaths attributable to days with area-wide concentrations of 60 

ppb or greater reflects the relatively small proportion of days with such elevated area-wide concentrations. 
4-40 

Figure 4-10. Estimates of O3-Associated Deaths Attributable to Full Distribution of 8-Hour Area-Wide O3 

to Concentrations at or above 20, 40, or 60 ppb - Deaths Summed Across Urban Case Study A

24As discussed in section 4.3.3 of the HREA, the model-based air quality adjustment approach used to estimate risks associated with the 

standards was unable to estimate the distribution of ambient O3 concentrations in New York City upon just meeting an alternative stand

Therefore, the total number of deaths indicated for the 60 ppb standard level in Figure 4-10 reflects the 60 ppb estimates for all urban cNew York City. For New York City, the estimated number of O3-associated deaths for the 65 ppb standard level was assumed. 
4-41 

Figure 4-11. Estimates of Respiratory Hospital Admissions Associated with Short-Term O3

Concentrations in Urban Case Study Areas (Air Quality Adjusted to Current 

and Potential alternative standard levels) – Total Risk 

2007 Simulation year

2009 Simulation year

4

6

8

10

12

14

16

75ppb 70ppb 65ppb 60ppb

Total ozone‐related HA per 100,000

Trend in ozone‐related HA across standard levels

(HA per 100,000)

Atlanta, GA

Baltimore, MD

Boston, MA

Cleveland, OH

Denver, CO

Detroit, MI

Houston, TX

Los Angeles, CA

New York, NY

Philadelphia, PA

Sacramento, CA

St. Louis, MO

4

6

8

10

12

14

16

75ppb 70ppb 65ppb 60ppb

Total ozone‐related HA per 100,000

Trend in ozone‐related HA across standard levels

(HA per 100,000)

Atlanta, GA

Baltimore, MD

Boston, MA

Cleveland, OH

Denver, CO

Detroit, MI

Houston, TX

Los Angeles, CA

New York, NY

Philadelphia, PA

Sacramento, CA

St. Louis, MO
4-42 

Key results from the HREA (U.S. EPA, 2014, chapter 7) are summarized in Figure 4-12 

below for estimates of respiratory mortality associated with long-term O3 concentrations, based 

on the study by Jerrett et al. (2009). As discussed in section 3.2.3.2 of this PA, Jerrett et al. 

(2009) reported that when seasonal averages of 1-hour daily maximum O3 concentrations ranged 

from 33 to 104 ppb, there was no statistical deviation from a linear concentration-response 

relationship between O3 and respiratory mortality across 96 U.S. cities (U.S. EPA, 2013, section 

7.7). However, the authors reported “limited evidence” for an effect threshold at an O3 

concentration of 56 ppb (p=0.06). In communications with EPA staff (described in Sasser, 2014), 

the study authors indicated that it is not clear whether a threshold model is a better predictor of 

respiratory mortality than the linear model, and that “considerable caution should be exercised in 

accepting any specific threshold.” Consistent with this communication, the HREA estimated 

respiratory mortality associated with long-term O3 concentrations based on the linear model from 

the published study, and in a series of sensitivity analyses with models that included thresholds 

ranging from 40 to 60 ppb (U.S. EPA, 2014, Figure 7-9). 

Figure 4-12 presents estimates of total O3-associated respiratory deaths, based on a linear 

concentration-response relationship. As discussed for the current standard (section 3.2.3.2), 

HREA sensitivity analyses indicate that, if a threshold exists between 40 and 60 ppb, the number 

of respiratory deaths associated with long-term O3 concentrations could potentially be 

considerably smaller than indicated by the no threshold model (U.S. EPA, 2014, Figure 7-9). 
4-43 

Figure 4-12. Estimates of Respiratory Mortality Associated with long-term O3

Concentrations in Urban Case Study Areas (Air Quality Adjusted to Current 

and Potential alternative standard levels) – Total Risk 

Based on Figures 4-9 to 4-12 and the associated details described in the HREA (U.S. 

EPA 2014, chapter 7), we take note of the following for an O3 standard level of 70 ppb: 

1. Total mortality associated with short-term O3 concentrations: 
4-44 

a. Across urban case study areas, risks are estimated to decrease by up to 

approximately 5% for a standard level of 70 ppb, compared to the current 

standard. Risk reductions are estimated consistently for the model year with 

generally higher O3-associated risks (2007). In the year with generally lower risks 

(2009), a standard level of 70 ppb results in either no change or more modest 

reductions in estimated risks in most urban case study areas. In one area (Detroit) 

for the 2009 model year, O3-associated mortality is estimated to increase by 

approximately 4%, compared to the current standard (see section 3.2.3.2 for 

further discussion of increased risk estimates following air quality adjustment25). 

b. When summed across urban case study areas, a standard level of 70 ppb is 

estimated to reduce O3-associated deaths by approximately 4% (2007 model year) 

and 2% (2009 model year), compared to the current standard. For area-wide 

concentrations at or above 40 ppb, a standard level of 70 ppb is estimated to 

reduce O3-associated deaths by approximately 10% (2007 model year) and 9% 

(2009 model year). For area-wide concentrations at or above 60 ppb, a standard 

level of 70 ppb is estimated to reduce O3-associated deaths by approximately 50% 

(2007 model year) and 70% (2009 model year).26 

2. Respiratory hospital admissions associated with short-term O3 concentrations: Compared to 

the current standard, changes in total risk estimated for a standard level of 70 ppb are similar 

to the changes in total risks estimated for total mortality (U.S. EPA, 2014, chapter 7). 

3. Respiratory mortality associated with long-term O3 concentrations: A standard level of 70 

ppb reduces total risk, compared to the current standard. Across urban case study areas, risks 

are estimated to decrease by up to approximately 6%. These risk reductions are estimated 

most consistently for the model year with generally higher O3-associated risks (2007). In the 

year with generally lower O3 concentrations (2009), a standard level of 70 ppb results in 

smaller reductions in estimated risks in most urban case study areas. In one area (Detroit) for 

the 2009 model year, O3-associated mortality is estimated to increase by approximately 1%, 

compared to the current standard. 

Based on Figures 4-9 to 4-12 and the associated details described in the HREA (U.S. 

EPA 2014, chapter 7), we take note of the following for an O3 standard level of 65 ppb: 

1. Total mortality associated with short-term O3 concentrations: 

a. Across most urban case study areas, risks are estimated to decrease by up to 

approximately 9% for a standard level of 65 ppb, compared to the current 

 

25As discussed in more detail above (section 3.2.3.2), because of the influence of the entire distribution of ambient 

O3 concentrations on total risk estimates, the impacts of adjusting air quality to just meet potential alternative 

standards are more modest, and are less directionally consistent across urban case study areas, than observed for 

exposures of concern or O3-induced lung function decrements. 

26These results reflect the fact that increases in area-wide O3 concentrations upon air quality adjustment occur 

primarily at relatively low concentrations (i.e., on days with initial O3 concentrations in the range of 10 to 40) (U.S. 

EPA, 2014, section 4.3.3.2 and appendix 7B, section 9.6). 
4-45 

standard. In one area (New York City), risks are estimated to decrease by up to 

approximately 22%.27 These risk reductions are estimated most consistently for 

the model year with generally higher O3-associated risks (2007). In the year with 

generally lower risks (2009), a standard level of 65 ppb results in smaller 

reductions in estimated risks in most urban case study areas. In one area (Detroit) 

for the 2009 model year, O3-associated mortality is estimated to increase by 

approximately 1% compared to the current standard. 

b. When summed across urban case study areas, a standard level of 65 ppb is 

estimated to reduce O3-associated deaths by approximately 13% (2007 model 

year) and 9% (2009 model year), compared to the current standard. For area-wide 

concentrations at or above 40 ppb, a standard level of 65 ppb is estimated to 

reduce O3-associated deaths by approximately 47% (2007) and 46% (2009). For 

area-wide concentrations at or above 60 ppb, a standard level of 65 ppb is 

estimated to reduce O3-associated deaths by over 80% (2007 and 2009 model 

years). 

2. Respiratory hospital admissions associated with short-term O3 concentrations: Compared to 

the current standard, changes in total risk estimated for a standard level of 65 ppb are similar 

to the changes in total risk estimated for total mortality (U.S. EPA, 2014, chapter 7). 

3. Respiratory mortality associated with long-term O3 concentrations: A standard level of 65 

ppb reduces total risk, compared to the current standard. Across most urban case study areas, 

risks are estimated to decrease by up to approximately 10%. In one area (New York City), 

risks are estimated to decrease by up to approximately 24%. Risk reductions are estimated 

across all urban case study areas and in both model years evaluated, with larger reductions 

estimated for 2007 (i.e., the model year with generally higher O3-associated risks). 

Based on Figures 4-9 to 4-12 and the associated details described in the HREA (U.S. 

EPA 2014, chapter 7), we take note of the following for an O3 standard level of 60 ppb: 

1. Total mortality associated with short-term O3 concentrations: 

a. A standard level of 60 ppb is estimated to reduce total risk, compared to the 

current standard, in all urban case study areas. Across urban case study areas, 

risks are estimated to decrease by up to approximately 14%. Estimated risk 

reductions are larger for the model year with generally higher O3-associated risks 

(2007). 

b. When summed across urban case study areas, a standard level of 60 ppb is 

estimated to reduce O3-associated deaths by approximately 15% (2007 model 

year) and 11% (2009 model year), compared to the current standard. For areawide concentrations at or above 40 ppb, a standard level of 60 ppb is estimated to 

 

27 Because of the approach to adjusting air quality in New York (and Los Angeles), which differed from other urban 

case study areas (U.S. EPA, 2014, sections 4.3.3.1, 4.5), the HREA notes less overall confidence in results for these 

areas. 
4-46 

reduce O3-associated deaths by almost 60% (2007 and 2009 model years). For 

area-wide concentrations at or above 60 ppb, a standard level of 60 ppb is 

estimated to reduce O3-associated deaths by over 95% (2007 and 2009 model 

years). 

2. Respiratory hospital admissions associated with short-term O3 concentrations: Compared to 

the current standard, changes in total risk estimated for a standard level of 60 ppb are similar 

to the changes in total risk estimated for total mortality (U.S. EPA, 2014, chapter 7). 

3. Respiratory mortality associated with long-term O3 concentrations: A standard level of 60 

ppb reduces total risk, compared to the current standard. Across urban case study areas, risks 

are estimated to decrease by up to approximately 17%. Risk reductions are estimated across 

all urban case study areas and in both model years evaluated, with larger reductions 

estimated for 2007 (i.e., the model year with generally higher O3-associated risks). 

In further considering these risk estimates, we take note of the associated uncertainties, as 

discussed in more detail in section 3.2.3.2 of this PA. In particular, these include (1) the national 

representativeness of urban case study areas in terms of the O3 response to reductions in NOX

emissions; (2) the representativeness of risk changes based primarily on reductions in NOX

emissions versus changes that could be achieved with better-optimized emissions reduction 

strategies; (3) the shape of the concentration-response function at lower ambient concentrations, 

including the potential for a threshold in the association between long-term O3 and respiratory 

mortality; (4) the presence of unexplained heterogeneity in effect estimates between locations; 

(5) the potential for exposure measurement errors; and (6) the possibility for reductions in risk 

associated with reductions in PM and/or NO2 resulting from control of NOX. 

4.5 CASAC ADVICE AND PUBLIC COMMENTERS’ VIEWS ON 

ALTERNATIVE STANDARDS 

As discussed in section 3.3, staff recognizes that decisions regarding the weight to place 

on various types of evidence, exposure/risk information, and associated uncertainties reflect 

public health policy judgments that are ultimately left to the Administrator. To help inform those 

judgments with regard to the range of alternative primary O3 standards appropriate for 

consideration, CASAC has provided advice to the Administrator based on their reviews of the O3

ISA, HREA, and PA. This section summarizes the advice provided by CASAC regarding 

potential alternative standards, as well as the views expressed at the CASAC meetings by public 

commenters. 

In the fall of 2011, rather than revising the O3 NAAQS as part of the reconsideration 

process, EPA elected to coordinate further proceedings on the reconsideration rulemaking with 

the current ongoing periodic review. Accordingly, in this section we briefly describe CASAC 

advice from the reconsideration of the 2008 final decision on the level of the standard, as well as 
4-47 

CASAC advice received during the current review as it pertains to potential alternative 

standards. 

Consistent with their advice in 2008, CASAC reiterated during the reconsideration its 

support for an 8-hour primary O3 standard with a level ranging from 60 to 70 ppb, combined 

with the current indicator, averaging time, and form. Specifically, in response to EPA’s 

solicitation of their advice during the reconsideration, the CASAC letter (Samet 2010) to the 

Administrator stated: 

CASAC fully supports EPA’s proposed range of 0.060 – 0.070 parts per million 

(ppm) for the 8-hour primary ozone standard. CASAC considers this range to be 

justified by the scientific evidence as presented in the Air Quality Criteria for 

Ozone and Related Photochemical Oxidants (March 2006) and Review of the 

National Ambient Air Quality Standards for Ozone: Policy Assessment of 

Scientific and Technical Information, OAQPS Staff Paper (July 2007). 

Similarly, in response to EPA’s request for additional advice on the reconsideration in 

2011, CASAC reaffirmed their conclusion that “the evidence from controlled human and 

epidemiological studies strongly supports the selection of a new primary ozone standard within 

the 60 – 70 ppb range for an 8-hour averaging time” (Samet, 2011). CASAC further concluded 

that this range “would provide little margin of safety at its upper end” (Samet, 2011, p. 2). 

In the current review of the Second Draft PA, as noted above, CASAC concurred with 

staff’s conclusions that it is appropriate to consider retaining the current indicator (O3), averaging 

time (8-hour average) and form (three-year average of the 4th highest maximum daily 8-hour 

average. With regard to level, CASAC stated the following (Frey, 2014, p. ii to iii): 

The CASAC further concludes that there is adequate scientific evidence to 

recommend a range of levels for a revised primary ozone standard from 70 ppb to 

60 ppb. The CASAC reached this conclusion based on the scientific evidence from 

clinical studies, epidemiologic studies, and animal toxicology studies, as 

summarized in the Integrated Science Assessment (ISA), the findings from the 

exposure and risk assessments as summarized in the HREA, and the interpretation 

of the implications of these sources of information as given in the Second Draft 

PA. 

The CASAC acknowledges that the choice of a level within the range 

recommended based on scientific evidence [i.e., 70 to 60 ppb] is a policy 

judgment under the statutory mandate of the Clean Air Act. The CASAC advises 

that, based on the scientific evidence, a level of 70 ppb provides little margin of 

safety for the protection of public health, particularly for sensitive 

subpopulations. 

Thus, our policy advice is to set the level of the standard lower than 70 ppb within 

a range down to 60 ppb, taking into account your judgment regarding the desired 
4-48 

margin of safety to protect public health, and taking into account that lower levels 

will provide incrementally greater margins of safety. 

 The public commenters who expressed the view that the current primary O3 standard is 

not adequate (section 3.3) also submitted comments that supported revising the level of the 

primary O3 standard. Several of these commenters expressed the view that the level should be 

revised to the lower end of the range of 70 to 60 ppb, or in some cases to a level below 60 ppb. 

The basis for these commenters’ views on the level of the standard is generally reflected in the 

rationale given by CASAC for their advice, and is discussed in section 3.3 of this PA. Public 

commenters who expressed the view that revision of the current standard is not necessary did not 

provide any provisional views on alternative levels that would be appropriate for consideration 

should the Administrator consider revisions to the standard. These views are also discussed in 

section 3.3 of this PA. 

4.6 STAFF CONCLUSIONS ON ALTERNATIVE PRIMARY STANDARDS FOR 

CONSIDERATION 

Staff’s consideration of alternative primary O3 standards builds upon our conclusion, 

discussed in section 3.4, that the overall body of evidence and exposure/risk information call into 

question the adequacy of public health protection afforded by the current standard, particularly 

for at-risk populations. We further conclude that it is appropriate in this review to consider 

alternative standards that would increase public health protection, compared to the current 

standard. 

As discussed in sections 4.1 to 4.3 above, in the current review we conclude that it is 

appropriate for the Administrator to consider retaining O3 as the indicator for the standard that 

protects against exposures to ambient O3 and other photochemical oxidants (section 4.1), and to 

consider retaining the current averaging time (section 4.2) and form (section 4.3) for the primary 

O3 standard. For a primary O3 standard that is defined in terms of the current indicator, averaging 

time, and form, we reach the conclusion that, depending on the public health policy judgments 

made by the Administrator, the scientific evidence and exposure/risk information available in 

this review support considering alternative O3 standard levels from 70 down to 60 ppb. The basis 

for this conclusion is discussed in detail in section 4.4 of this PA, and is summarized in this 

section. 

Below, we summarize our approach to considering the scientific evidence and 

exposure/risk information, and the specific evidence and information that supports the range of 

levels from 70 to 60 ppb. In doing so, we focus particularly on the evidence and information as it 

relates to the upper (70 ppb), middle (65 ppb), and lower (60 ppb) portions of this range. Key 

exposure/risk information is summarized in Tables 4-4, and 4-5, and Figure 4-13. 
4-49 

Table 4-4 Summary of Estimated Exposures of Concern for Potential Alternative O3

Standard Levels of 70, 65, and 60 ppb in Urban Case Study Areas28

Benchmark 

Level 

Alternative 

Standard 

Level (ppb) 

Average % 

Children 

Exposed29

Number of Children (5 

to 18 years) 

[Number of Asthmatic 

Children]30

Average % 

Reduction from 

Current 

Standard 

% Children - 

Worst Year and 

Worst Area 

One or more exposures of concern per season 

≥ 70 ppb 

70 0.1-1.2 94,000 [10,000] 73 3.2 

65 0-0.2 14,000 [2,000] 95 0.5 

60 031 1,400 [200]32 100 0.1 

≥ 60 ppb 

70 3.3-10.2 1,176,000 [126,000] 46 18.9 

65 0-4.2 392,000 [42,000] 80 9.5 

60 0-1.2 70,000 [8,000] 96 2.2 

Two or more exposures of concern per season 

≥ 70 ppb 

70 0-0.1 5,400 [600] 95 0.4 

65 0 300 [100] 100 0 

60 0 0 [0] 100 0 

≥ 60 ppb 

70 0.5-3.5 320,000 [35,000] 61 9.2 

65 0-0.8 67,000 [7,500] 92 2.8 

60 0-0.2 5,100 [700] 100 0.3 

 

28 As illustrated above in Figures 4-1 to 4-4, all alternative standard level s evaluated in the HREA were effective at 

limiting exposures of concern at or above 80 ppb. Therefore, Table 4-4 focuses on exposures of concern at or above 

the 70 and 60 ppb benchmark concentrations. 

29 Estimates for each urban case study area were averaged for the years evaluated in the HREA (2006 to 2010). 

Ranges reflect the ranges across urban case study areas. 

30 Numbers of children exposed in each urban case study area were averaged over the years 2006 to 2010. These 

averages were then summed across urban case study areas. Numbers are rounded to nearest thousand unless 

otherwise indicated. 

31 Estimates smaller than 0.1% were rounded to zero. 32As discussed in section 4.3.3 of the HREA, the model-based air quality adjustment approach used to estimate risks 

associated with the current and alternative standards was unable to estimate the distribution of ambient O3

concentrations in New York City upon just meeting an alternative standard with a level of 60 ppb. Therefore, for the 

60 ppb standard level the numbers of children and asthmatic children reflect all of the urban case study areas except 

New York. 
4-50 

Table 4-5 Summary of Estimated Lung Function Decrements for Potential Alternative 

O3 Standard Levels of 70, 65, and 60 ppb in Urban Case Study Areas 

Lung 

Function 

Decrement 

Alternative 

Standard 

Level 

Average % 

Children33

Number of Children (5 

to 18 years) [Number of 

Asthmatic Children]34

Average % 

Reduction from 

Current Standard 

% Children 

Worst Year and 

Area 

One or more decrements per season 

≥ 10% 

70 11-17 2,527,000 [261,000] 15 20 

65 3-15 1,896,000 [191,000] 31 18 

60 5-11 1,404,000 [139,000]35 45 13 

≥ 15% 

70 2-4 562,000 [58,000] 26 5 

65 0-3 356,000 [36,000] 50 4 

60 1-2 225,000 [22,000] 67 3 

≥ 20% 

70 1-2 189,000 [20,000] 32 2.1 

65 0-1 106,000 [11,000] 59 1.4 

60 0-1 57,000 [6,000] 77 0.7 

Two or more decrements per season 

≥ 10% 

70 5.5-11 1,414,000 [145,000] 17 13 

65 1.3-8.8 1,023,000 [102,000] 37 11 

60 2.1-6.4 741,000 [73,000] 51 7.3 

≥ 15% 

70 0.9-2.4 276,000 [28,000] 29 3.1 

65 0.1-1.8 168,000 [17,000] 54 2.3 

60 0.2-1.0 101,000 [10,000] 71 1.4 

≥ 20% 

70 0.3-0.8 81,000 [8,000] 34 1.1 

65 0-0.5 43,000 [4,000] 66 0.8 

60 0-0.2 21,000 [2,000] 83 0.4 

 

33 Estimates in each urban case study area were averaged for the years evaluated in the HREA (2006 to 2010). 

Ranges reflect the ranges across urban case study areas. 

34 Numbers of children estimated to experience decrements in each study urban case study area were averaged over 

2006 to 2010. These averages were then summed across urban case study areas. Numbers are rounded to nearest 

thousand unless otherwise indicated. As discussed above, for the 60 ppb standard level the numbers of children and 

asthmatic children included in Table 4-5 reflect all of the urban case study areas except New York. 35As discussed in section 4.3.3 of the HREA, the model-based air quality adjustment approach used to estimate risks 

associated with the current and alternative standards was unable to estimate the distribution of ambient O3

concentrations in New York City upon just meeting an alternative standard with a level of 60 ppb. Therefore, for the 

60 ppb standard level the numbers of children and asthmatic children reflect all of the urban case study areas except 

New York. 
4-51 

Figure 4-13. Estimates of O3-Associated Deaths Attributable to Full Distributions of 8-

Hour Area-Wide O3 Concentrations and to Concentrations at or above 20, 40, or 6036 ppb 

O3 - Deaths Summed Across Urban Case Study Areas and Expressed Relative to a 

Standard with a Level of 75 ppb 

 

 

36As discussed in section 4.3.3 of the HREA, the model-based air quality adjustment approach used to estimate risks 

associated with the current and alternative standards was unable to estimate the distribution of ambient O3

concentrations in New York City upon just meeting an alternative standard with a level of 60 ppb. Therefore, the 

total number of deaths indicated for the 60 ppb standard level in Figure 4-10 reflects the 60 ppb estimates for all 

urban case study areas except New York City. For New York City, the estimated number of O3-associated deaths for 

the 65 ppb standard level was assumed. 
4-52 

Summary of approach to reaching conclusions on alternative standard levels 

In this PA, our approach to reaching conclusions on alternative standard levels focuses on 

the evidence from controlled human exposure and epidemiologic studies, as assessed in the ISA 

(U.S. EPA, 2013), and the exposure and health risk analyses presented in the HREA (U.S. EPA, 

2014). This approach is discussed in detail in Chapter 1 (section 1.3), and is summarized below. 

As an initial matter, we note that controlled human exposure studies provide the most 

certain evidence indicating the occurrence of health effects in humans following exposures to 

specific O3 concentrations. Consistent with this, CASAC concluded that “the scientific evidence 

supporting the finding that the current standard is inadequate to protect public health is strongest 

based on the controlled human exposure studies of respiratory effects” (Frey, 2014, p. 5). As 

discussed above and in section 3.1.2.1, controlled human exposure studies have reported a 

variety of respiratory effects in healthy adults following exposures to O3 concentrations of 60, 

72,37 or 80 ppb, and higher. The largest respiratory effects, and the broadest range of effects, 

have been studied and reported following exposures of healthy adults to 80 ppb O3 or higher, 

with most exposure studies conducted at these higher concentrations. Exposures to O3

concentrations of 80 ppb or higher have been reported to decrease lung function, increase airway 

inflammation, increase respiratory symptoms, result in airway hyperresponsiveness, and decrease 

lung host defenses in healthy adults. 

Most of these effects have also been reported in healthy adults following exposures to O3

concentrations below 80 ppb.38 Exposures to O3 concentrations of 72 ppb have been reported to 

decrease lung function and increase respiratory symptoms, a combination that meets the ATS 

criteria for an “adverse” response (section 3.1.3). Exposures to O3 concentrations of 60 ppb have 

been demonstrated to decrease lung function, with decrements in some people large enough to be 

judged an abnormal response by ATS, and which CASAC has indicated could be adverse to 

people with lung disease.39 In addition, as discussed in section 3.1.3, such a decrease in mean 

lung function meets the ATS criteria for an adverse response given that a downward shift in the 

distribution of FEV1 would result in diminished reserve function, and therefore would increase 

risk from further environmental insult. Exposures to O3 concentrations of 60 ppb have also been 

reported in one study (Kim et al., 2011) to increase airway inflammation, which provides a 

 

37 As noted above, for the 70 ppb exposure concentration Schelegle et al. (2009) reported that the actual mean 

exposure concentration was 72 ppb. 

38 Airway hyperresponsiveness and reductions in lung host defense have not been evaluated following exposures to 

O3 concentrations below 80 ppb. The extent to which these respiratory effects occur following lower exposure 

concentrations is not clear from the available evidence, though we have no basis for concluding that an exposure 

concentration of 80 ppb reflects an effects threshold. 

39 In their advice to the Administrator based on the second draft PA, the CASAC indicated that “60 ppb is an 

appropriate exposure of concern for asthmatic children” (Frey, 2014). 
4-53 

mechanism by which O3 may cause other more serious respiratory effects (e.g., asthma 

exacerbations). 

Given the evidence for respiratory effects from controlled human exposure studies, we 

consider the extent to which standards with revised levels would be estimated to protect at-risk 

populations against exposures of concern to O3 concentrations at or above the health benchmark 

concentrations of 60, 70, and 80 ppb (i.e., based on HREA estimates of one or more and two or 

more exposures of concern). In doing so, we note that, due to individual variability in 

responsiveness, only a subset of people who experience exposures at or above the three 

benchmark concentrations can be expected to experience associated health effects, and that 

available data are not sufficient to quantify that subset of people. We view the health effects 

evidence as a continuum with greater confidence and less uncertainty about the occurrence of 

adverse health effects at higher O3 exposure concentrations, and less confidence and greater 

uncertainty as one considers lower exposure concentrations (discussed in more detail in section 

3.2.2). 

While there is greater uncertainty regarding the occurrence of adverse health effects at 

lower concentrations, we also note that the controlled human exposure studies that provided the 

basis for benchmark concentrations have not evaluated responses in populations at the greatest 

risk from exposures to O3. Thus, the effects reported in healthy adults at each of the benchmark 

concentrations may underestimate effects in these at-risk groups. Compared to the healthy people 

included in most controlled human exposure studies, members of at-risk populations, including 

lifestages, (e.g., asthmatics, children) are at greater risk of experiencing adverse effects. In 

considering the health evidence within the context of drawing conclusions on potential 

alternative standard levels, we balance concerns about the potential for adverse health effects, 

especially in at-risk populations, with our increasing uncertainty regarding the likelihood of such 

effects following exposures to lower O3 concentrations. 

With respect to the lung function decrements that have been evaluated in controlled 

human exposure studies, we consider the extent to which standards with revised levels would be 

estimated to protect healthy and at-risk populations against O3-induced lung function decrements 

large enough to be adverse in some people (based on quantitative risk estimates in the HREA). 

As discussed in section 3.1.3, although some experts would judge single occurrences of moderate 

responses to be a “nuisance,” especially for healthy individuals, a more general consensus view 

of the adversity of moderate lung function decrements emerges as the frequency of occurrence 

increases. Repeated occurrences of moderate responses, even in otherwise healthy individuals, 

may be considered to be adverse, since they could well set the stage for more serious illness (61 

FR 65723). For the purpose of estimating potentially adverse lung function decrements in active, 

healthy people, in the 2008 review the CASAC panel indicated that a focus on the mid to upper 
4-54 

end of the range of moderate (i.e., FEV1 decrements ≥ 15%) functional responses is appropriate. 

However, for children and adults with lung disease, FEV1 decrements ≥ 10% could lead to 

respiratory symptoms, would likely interfere with normal activities for many individuals, and 

therefore could be adverse. Large (i.e., FEV1 decrements ≥ 20%) lung function decrements 

would likely interfere with normal activities for most people with lung disease and would 

increase the likelihood that they would seek medical attention. In the current review, CASAC 

judges that an FEV1 decrement ≥ 15% is an appropriate surrogate for adverse health outcomes in 

active healthy adults, while a decrement ≥ 10% is a scientifically relevant surrogate for adverse 

health outcomes for people with asthma and lung disease (Frey, 2014). In reaching conclusions 

on alternative standard levels, we consider the extent to which standards with revised levels 

would be estimated to protect healthy and at-risk populations against one or more, and two or 

more, moderate (i.e., FEV1 decrements ≥ 10% and ≥ 15%) and large (i.e., FEV1 decrements ≥

20%) lung function decrements. 

In evaluating the epidemiologic evidence within the context of drawing conclusions on 

potential alternative standard levels, we consider the extent to which available studies have 

reported associations with emergency department visits, hospital admissions, and/or mortality in 

locations that would likely have met potential alternative standards with levels below 75 ppb 

(based on analyses presented in section 4.4.1). In evaluating the epidemiologic evidence in this 

way, we consider both multicity and single-city studies, recognizing the strengths and limitations 

of each. Specifically, multicity studies evaluate large populations and provide greater statistical 

power than single-city studies; multicity studies reflect O3-associated health impacts across a 

range of diverse locations, providing spatial coverage for different regions across the country and 

reflecting differences in exposure-related factors that could impact O3 risks; and multicity studies 

afford a greater possibility of generalizing to the national population. In contrast, while singlecity studies are more limited than multicity studies in terms of statistical power and geographic 

coverage, conclusions linking air quality in a specific area with health effect associations in that 

same area can be made with greater certainty for single-city studies (i.e., compared to multicity 

studies reporting only multicity effect estimates). 

We also consider the epidemiologic evidence within the context of epidemiology-based 

risk estimates. Compared to the weight given to HREA estimates of exposures of concern and 

lung function risks (sections 4.4.2.1 and 4.4.2.2, above), and the weight given to the evidence 

(section 4.4.1), we place relatively less weight on epidemiologic-based risk estimates. In doing 

so, we note that the overall conclusions from the HREA likewise reflect less confidence in 

estimates of epidemiologic-based risks than in estimates of exposures and lung function risks. 

Our determination to attach less weight to the epidemiologic-based estimates reflects the 

uncertainties associated with mortality and morbidity risk estimates, including the heterogeneity 
4-55 

in effect estimates between locations, the potential for exposure measurement errors, and 

uncertainty in the interpretation of the shape of concentration-response functions at lower O3

concentrations (U.S. EPA, 2014, section 9.6). The HREA also concludes that lower confidence 

should be placed in the results of the assessment of respiratory mortality risks associated with 

long-term O3 exposures, primarily because that analysis is based on only one study (even though 

that study is well-designed) and because of the uncertainty in that study about the existence and 

level of a potential threshold in the concentration-response function (U.S. EPA, 2014, section 

9.6). 

In considering the epidemiology-based risk estimates, we focus on the extent to which 

potential alternative O3 standards with levels below 75 ppb are estimated to reduce the risk of 

O3-associated mortality (based on the HREA results summarized in section 4.4.2.3).40 As 

discussed in section 3.4 for the current standard, we consider estimates of total risk (i.e., based 

on the full distributions of ambient O3 concentrations) and estimates of risk associated with O3

concentrations in the upper portions of ambient distributions. A focus on estimates of total risks 

would place greater weight on the possibility that concentration-response relationships remain 

linear over the entire distribution of ambient O3 concentrations, and thus on the potential for 

mortality and morbidity to be affected by changes in relatively low O3 concentrations. A focus 

on risks associated with O3 concentrations in the upper portions of the ambient distribution 

would place greater weight on the uncertainty associated with the shapes of concentrationresponse curves for O3 concentrations in the lower portions of the distribution. Given that both 

types of risk estimates could reasonably inform a decision on standard level, depending on the 

weight placed on uncertainties in the occurrence and the estimation of O3-attributable effects at 

relatively low O3 concentrations, in reaching conclusions we consider what both types of 

estimates indicate with regard to potential alternative levels. 

Staff conclusions on the range of levels appropriate for consideration 

Using the approach discussed above to consider the scientific evidence and exposure/risk 

information, we reach the conclusion that it is appropriate for the Administrator to consider 

alternative primary O3 standard levels from 70 to 60 ppb. The basis for this conclusion is 

discussed in detail in sections 4.4.1 and 4.4.2 above, and is summarized below. 

With regard to controlled human exposure studies, we consider the lowest O3 exposure 

concentrations at which various effects have been evaluated and statistically significant effects 

reported. We also consider the potential for reported effects to be adverse, including in at-risk 

populations and lifestages. As discussed in section 3.1.2.1, controlled human exposure studies 

 

40 Differences in estimated respiratory morbidity risks between potential alternative standard levels are similar to the 

differences estimated for total mortality associated with short-term O3 concentrations. 
4-56 

provide evidence of respiratory symptoms combined with lung function decrements (an 

“adverse” response based on ATS criteria) in healthy adults following exposures to O3

concentrations as low as 72 ppb, and evidence of potentially adverse lung function decrements 

and airway inflammation following exposures to O3 concentrations as low as 60 ppb. Although 

some studies show that respiratory symptoms also develop during exposures to 60 ppb O3, the 

increase in symptoms has not been reported to reach statistical significance by the end of the 6.6 

hour exposure period (Adams 2006; Schelegle et al. 2009). Thus, while significant increases in 

respiratory symptoms combined with lung function decrements have not been reported following 

exposures to 60 ppb O3, this combination of effects is likely to occur to some degree in healthy 

adults with 6.6-hour exposures to concentrations below 72 ppb, and also are more likely to occur 

with longer (i.e., 8-hour) exposures.41 

With regard to the lowest exposure concentration shown to cause respiratory effects (i.e., 

60 ppb), we note that most controlled human exposure studies have not evaluated O3

concentrations below 60 ppb. Therefore, 60 ppb does not necessarily reflect an exposure 

concentration below which effects such as lung function decrements and airway inflammation no 

longer occur. This is particularly the case given that controlled human exposure studies were 

conducted in healthy adults, while people with asthma, including asthmatic children, are likely to 

be more sensitive to O3-induced respiratory effects. In support of this, some epidemiologic panel 

studies, which can include members of at-risk groups such as children and outdoor workers, have 

found respiratory effects at ambient concentrations lower than 60 ppb (section 3.1.2.1). 

With regard to other O3-induced effects, we note that airway hyperresponsiveness and 

impaired lung host defense capabilities have been reported in healthy adults engaged in moderate 

exertion following exposures to O3 concentrations as low as 80 ppb, the lowest concentration 

evaluated for these effects. As discussed in section 3.1.2.1, these physiological effects have been 

linked to aggravation of asthma and increased susceptibility to respiratory infection, potentially 

leading to increased medication use, increased school and work absences, increased visits to 

doctors’ offices and emergency departments, and increased hospital admissions. These are all 

indicators of adverse O3-related morbidity effects, which are consistent with, and provide 

plausibility for, the adverse morbidity effects and mortality effects observed in epidemiologic 

studies. 

 

41 In addition, CASAC observed that, “adverse health effects in young healthy adults occur with exposures to 72 ppb 

of ozone for 6.6 hours” and that “It is the judgment of CASAC that if subjects had been exposed to ozone using the 

8-hour averaging period used in the standard, adverse effects could have occurred at [a] lower concentration. 

Further, in our judgment, the level at which adverse effects might be observed would likely be lower for more 

sensitive subgroups, such as those with asthma” (Frey, 2014, p. 5). 
4-57 

Based on consideration of the above evidence, we conclude that available controlled 

human exposure studies support a level no higher than 70 ppb as the upper end of the range for 

consideration in the current review. In reaching this conclusion, we note that 70 ppb is just below 

the O3 exposure concentration reported to result in lung function decrements and respiratory 

symptoms in healthy adults (i.e., 72 ppb), a combination of effects that meet ATS criteria for an 

adverse response. In addition, while 70 ppb is well below the 80 ppb exposure concentration 

shown to cause potentially adverse respiratory effects such as airway hyperresponsiveness and 

impaired host-defense capabilities, these effects have not been evaluated at exposure 

concentrations below 80 ppb and there is no reason to believe that 80 ppb represents a threshold 

for such effects. 

We further conclude that the evidence from controlled human exposure studies42 supports 

considering alternative O3 standard levels at least as low as 60 ppb. Potentially adverse lung 

function decrements and pulmonary inflammation have been demonstrated to occur in healthy 

adults at 60 ppb. Thus, 60 ppb is a short-term exposure concentration that may be reasonably 

concluded to elicit adverse effects in at-risk groups. Pulmonary inflammation, particularly if 

experienced repeatedly, provides a mechanism by which O3 may cause other more serious 

respiratory morbidity effects (e.g., asthma exacerbations) and possibly extrapulmonary effects. 

As discussed in section 3.1.2.1, the physiological effects reported in controlled human exposure 

studies down to 60 ppb O3 have been linked to aggravation of asthma and increased 

susceptibility to respiratory infection, potentially leading to increased medication use, increased 

school and work absences, increased visits to doctors’ offices and emergency departments, and 

increased hospital admissions. 

We further note that the range of alternative levels from 70 to 60 ppb is supported by 

evidence from epidemiologic studies and by exposure and risk estimates from the HREA. This 

evidence and exposure/risk information indicate that a level from anywhere in the range of 70 to 

60 ppb would be expected to result in important public health improvements over the current 

standard. In particular, compared to the current standard a revised standard with a level from 70 

to 60 ppb would be expected to (1) more effectively maintain short- and long-term O3

concentrations below those present in the epidemiologic studies that reported significant O3

health effect associations in locations likely to have met the current standard; (2) reduce the 

occurrence of exposures of concern to O3 concentrations that result in respiratory effects in 

healthy adults (at or above 60, 70, and 80 ppb); (3) reduce the occurrence of moderate-to-large 

O3-induced lung function decrements; and (4) reduce the risk of O3-associated mortality and 

 

42 As discussed in sections 3.1.2.1 and 4.4.1above, panel studies also provide supporting evidence for these 

conclusions. 
4-58 

morbidity, particularly the risk associated with the upper portions of the distributions of ambient 

O3 concentrations. 

In reaching a conclusion on whether it is appropriate to consider alternative standard 

levels below 60 ppb, we note the following: 

 While controlled human exposure studies provide evidence for O3-induced respiratory 

effects following exposures to O3 concentrations as low as 60 ppb, they do not provide 

evidence for adverse effects following exposures to lower concentrations. On this issue, 

CASAC concurred that 60 ppb O3 is an appropriate and justifiable scientifically based 

lower bound for a revised primary standard, based upon findings of “adverse effects, 

including clinically significant lung function decrements and airway inflammation, after 

exposures to 60 ppb ozone in healthy adults with moderate exertion (Adams 2006; Schelegle 

et al., 2009; Brown et al. 2008; Kim et al., 2011), with limited evidence of adverse effects 

below 60 ppb” (Frey, 2014, p. 7). 

 Based on the HREA results, meeting an O3 standard with a level of 60 ppb would be 

expected to almost eliminate exposures of concern to O3 concentrations at or above 60 

ppb. To the extent lower exposure concentrations may result in adverse health effects in 

some people, a standard level of 60 ppb would be expected to also reduce exposures to 

O3 concentrations below 60 ppb. 

 U.S. and Canadian epidemiologic studies have not reported O3 health effect associations 

based primarily on study locations likely to have met a standard with a level of 60 ppb. 

 In all of the urban case study areas evaluated, a standard with a level of 60 ppb would be 

expected to maintain long-term O3 concentrations below those where a key study 

indicates the most confidence in a linear concentration-response relationship with 

respiratory mortality. 

Beyond the above considerations, we also note the HREA estimates indicating that 

meeting an O3 standard with a level of 60 ppb would result in important reductions in the risk of 

O3-induced lung function decrements and O3-associated mortality and morbidity. Although some 

risk is estimated to remain based on these metrics, even with a level of 60 ppb, we have 

decreasing confidence in further public health improvements with levels below 60 ppb. We reach 

this conclusion because, as noted above, at a level of 60 ppb virtually no one in the population 

would be expected to experience exposures to O3 concentrations at or above 60 ppb under 

conditions demonstrated in controlled human exposure studies to result in respiratory effects, and 

because epidemiologic studies have not reported O3 health effect associations based primarily on 

study locations likely to have met a standard with a level of 60 ppb. Given all of the above 

considerations we conclude that, compared to standards with levels from 70 to 60 ppb, the extent 

to which standards with levels below 60 ppb could result in further public health improvements 
4-59 

becomes notably less certain. Therefore, we conclude that it is not appropriate in this review to 

consider standard levels below 60 ppb. 

The range of levels from 70 to 60 ppb corresponds to the range of levels recommended 

for consideration by CASAC, based on the available evidence and information (Frey, 2014). 

While CASAC further offered the “policy advice” to set the level below 70 ppb, based on margin 

of safety considerations, the Committee acknowledged that “the choice of a level within the 

range recommended based on scientific evidence [i.e., 70 to 60 ppb] is a policy judgment under 

the statutory mandate of the Clean Air Act” (Frey, 2014). Therefore, we note that our 

conclusions on the appropriate range for alternative primary O3 standard levels are consistent 

with CASAC conclusions that the scientific evidence and exposure/risk information supports 

consideration of levels from 70 to 60 ppb, and that the ultimate identification of a standard that 

protects public health with an adequate margin of safety will reflect policy judgments that are 

explicitly reserved for the Administrator (section 1.2.1). 

The following sections summarize the specific scientific evidence and exposure/risk 

information as they relate to revised O3 standards with levels from the upper (70 ppb), middle 

(65 ppb), and lower (60 ppb) portions of the range of 70 to 60 ppb. 

O3 standard level of 70 ppb 

A level of 70 ppb is below 80 ppb, an O3 exposure concentration that has been reported to 

elicit a range of respiratory effects that includes airway hyperresponsiveness and decreased lung 

host defense, in addition to lung function decrements, airway inflammation, and respiratory 

symptoms. A level of 70 ppb is also below the lowest exposure concentration at which the 

combined occurrence of respiratory symptoms and lung function decrements have been reported 

(i.e., 72 ppb), a combination judged adverse by the ATS (section 3.1.3). A level of 70 ppb is 

above the lowest exposure concentration demonstrated to result in lung function decrements 

large enough to be judged an abnormal response by ATS and above the lowest exposure 

concentration demonstrated to result in pulmonary inflammation (i.e., 60 ppb). 

Compared to the current standard, the HREA estimates that a revised O3 standard with a 

level of 70 ppb would reduce exposures of concern to O3 concentrations of 60, 70, and 80 ppb in 

urban case study areas, with such a standard level estimated to be most effective at limiting 

exposures at or above the higher health benchmark concentrations and at limiting multiple 

occurrences of such exposures. On average over the years 2006 to 2010, a standard with a level 

of 70 ppb is estimated to allow only up to about 1% of children (i.e., ages 5 to 18) to experience 

exposures of concern at or above 70 ppb (73% reduction, compared to current standard), and far 

less than 1% to experience two or more such exposures (95% reduction, compared to current 

standard). In the worst-case location and year (i.e., location and year with the largest exposure 
4-60 

estimate), about 3% of children are estimated to experience one or more exposures of concern at 

or above 70 ppb, and less than 1% are estimated to experience two or more. A standard with a 

level of 70 ppb is estimated to allow far less than 1% of children to experience exposures of 

concern at or above the 80 ppb benchmark concentration, even in the worst-case year (Table 4-

4).43 

An O3 standard with a level of 70 ppb is estimated to allow about 3 to 10% of children, 

including asthmatic children, to experience one or more exposures of concern at or above 60 ppb 

in a single O3 season. As noted above, CASAC advised EPA that 60 ppb is an appropriate 

exposure of concern with respect to adverse effects on people with asthma, including children 

(Frey, 2014, p. 6, 8). Compared to the current standard, this reflects about a 46% reduction, on 

average across the urban case study areas. A standard with a level of 70 ppb is estimated to allow 

about 1% to 4% of children to experience two or more exposures of concern at or above 60 ppb. 

In the worst-case location and year, a standard set at 70 ppb is estimated to allow about 19% of 

children to experience one or more exposures of concern at or above 60 ppb, and 9% to 

experience two or more such exposures (Table 4-4). 

Compared to the current standard, the HREA estimates that a revised O3 standard with a 

level of 70 ppb would also reduce O3-induced lung function decrements in children. A level of 

70 ppb is estimated to be most effective at limiting the occurrences of moderate and large lung 

function decrements (i.e., FEV1 decrements ≥ 15% and ≥ 20%, respectively), and at limiting 

multiple occurrences of O3-induced decrements. On average over the years 2006 to 2010, a 

standard with a level of 70 ppb is estimated to allow about 2 to 4% of children in the urban case 

study areas to experience one or more moderate O3-induced lung function decrements (i.e., FEV1

decrement ≥ 15%), which would be of concern for healthy people, and about 1 to 2.5% of 

children to experience two or more such decrements (approximately 30% reduction, compared to 

the current standard). In the worst-case location and year, up to 5% of children are estimated to 

experience one or more O3-induced lung function decrements ≥ 15%, and up to 3% are estimated 

to experience two or more such decrements. A standard set at 70 ppb is estimated to allow about 

2% or fewer children to experience large O3-induced lung function decrements (i.e., FEV1

decrement ≥ 20%), and to allow about 1% or fewer children to experience two or more such 

decrements, even in the worst-case years and locations (Table 4-5). 

On average over the years 2006 to 2010, an O3 standard set at 70 ppb is estimated to 

allow about 11 to 17% of children in the urban case study areas to experience one or more 

moderate O3-induced lung function decrements (i.e., FEV1 decrement ≥ 10%), which could be 

 

43 As noted above, due to interindividual variability, children (or adults) exposed at these levels will not necessarily 

experience health effects; the information available for some health effects is not sufficient to quantify the numbers 

of children in the urban case study areas who might experience these effects. 
4-61 

adverse for people with lung disease. This reflects an average reduction of about 15%, compared 

to the current standard. A standard with a level of 70 ppb is also estimated to allow about 6 to 

11% of children to experience two or more such decrements (17% reduction, compared to 

current standard). In the worst-case location and year, a standard set at 70 ppb is estimated to 

allow about 20% of children in the urban case study areas to experience one or more O3-induced 

lung function decrements ≥ 10%, and 13% to experience two or more such decrements (Table 4-

5). 

Compared to the current standard, a revised standard with a level of 70 ppb would also 

more effectively maintain short-term ambient O3 concentrations below those present in the 

epidemiologic studies that reported significant O3 health effect associations in locations likely to 

have met the current standard. In particular, the single-city study by Mar and Koenig (2009) 

reported positive and statistically significant associations with respiratory emergency department 

visits in children and adults in a location that likely would have met the current O3 standard over 

the entire study period but violated a revised standard with a level of 70 ppb or below. Thus, 

none of the single-city studies evaluated in section 4.4.1 provide evidence for O3 health effect 

associations in locations meeting a standard with a level of 70 ppb or below. While this analysis 

does not provide information on the extent to which the reported O3-associated emergency 

department visits would persist upon meeting an O3 standard with a level of 70 ppb, or on the 

extent to which standard levels below 70 ppb could further reduce the incidence of such 

emergency department visits, it suggests that a revised O3 standard with a level at or below 70 

ppb would require reductions in the ambient O3 concentrations that provided the basis for the 

health effect associations reported by Mar and Koenig. 

As discussed above, compared to single-city studies, there is greater uncertainty in 

linking air quality concentrations from individual study cities to multicity effect estimates. With 

regard to multicity studies, we note that Dales et al. (2006) reported significant associations with 

respiratory hospital admissions based on air quality in 11 Canadian cities, most of which would 

likely have met the current standard over the entire study period but violated a revised standard 

with a level of 70 ppb or below over at least part of that period (Table 4-1). This analysis 

suggests that while the current standard would allow the ambient O3 concentrations in most of 

the study locations that provided the basis for the association with hospital admissions, a revised 

O3 standard with a level at or below 70 ppb would require reductions in those ambient O3

concentrations. As with the study by Mar and Koenig, this analysis does not provide information 

on the extent to which the reported O3-associated hospital admissions would persist upon 
4-62 

meeting an O3 standard with a level of 70 ppb, or on the extent to which standard levels below 70 

ppb could further reduce the incidence of such hospital admissions.44 

With regard to long-term O3 concentrations, we evaluated the “long-term” O3 metrics 

reported to be associated with mortality or morbidity in recent epidemiologic studies (e.g., 

seasonal averages of 1-hour or 8-hour daily max concentrations). Compared to the current 

standard, a revised standard with a level of 70 ppb would be expected to reduce the risk of 

respiratory mortality associated with long-term O3 concentrations, based on information from the 

study by Jerrett et al. (2009), though we note the HREA conclusion, discussed above, that lower 

confidence should be placed in respiratory mortality risk estimates based on this study (U.S. 

EPA, 2014, section 9.6). In addition, a standard with a level of 70 ppb would be expected to 

more effectively maintain long-term O3 concentrations below those where the study by Jerrett et 

al. (2009) indicates the most confidence in the reported association with respiratory mortality. 

Specifically, air quality analyses indicate this to be the case in 9 out of the 12 urban case study 

areas for a level of 70 ppb, compared to 6 out of 12 areas for the current standard. Finally, a 

revised standard with a level of 70 ppb would be expected to reduce long-term O3 concentrations 

based on the types of metrics that have been reported in recent epidemiologic studies to be 

associated with respiratory morbidity (i.e., seasonal averages of daily maximum 8-hour 

concentrations). 

In further considering the potential implications of epidemiology studies for alternative 

standard levels, we note estimates of total mortality associated with short-term O3

concentrations.45 As discussed above, we consider estimates of total risk (i.e., based on the full 

distributions of ambient O3 concentrations) and estimates of risk associated with O3

concentrations in the upper portions of ambient distributions. With regard to total risk we note 

that, when summed across urban case study areas, a standard with a level of 70 ppb is estimated 

to reduce the number of deaths associated with short-term O3 concentrations by about 4% (2007) 

and 2% (2009), compared to the current standard.46 Based on a national modeling analysis, the 

majority of the U.S. population would be expected to experience reductions in such risks upon 

reducing precursor emissions. 

 

44 In addition, for the other multicity studies identified in Table 4-1 (Cakmak et al., 2006; Stieb et al., 2009; 

Katsouyanni et al., 2009), and for the study by Bell et al. (2006) (for the 30 ppb cut point) (Table 4-2), the majority 

of study locations would likely have met a standard with a level of 70 ppb. 

45 As discussed above, compared to the weight given to the evidence and to HREA estimates of exposures of 

concern and lung function risks, we place relatively less weight on epidemiologic-based risk estimates. 46 A standard with a level of 70 ppb is also estimated to reduce respiratory mortality associated with long-term O3

concentrations in urban case study areas. However, given uncertainties associated with these risk estimates, as 

discussed above, we give them limited weight. 
4-63 

Compared to the total risk estimates noted above, an O3 standard with a level of 70 ppb is 

estimated to be more effective at reducing the number of deaths associated with short-term O3

concentrations at the upper ends of ambient distributions. Specifically, for area-wide O3

concentrations at or above 40 ppb, a standard with a level of 70 ppb is estimated to reduce the 

number of deaths associated with short-term O3 concentrations by about 10% compared to the 

current standard. In addition, for area-wide concentrations at or above 60 ppb, a standard with a 

level of 70 ppb is estimated to reduce O3-associated deaths by about 50% to 70% (Figure 4-13). 

As discussed above, CASAC concluded that there is adequate scientific evidence to 

consider a range of levels for a primary standard that includes an upper end at 70 ppb. However, 

CASAC differentiated its advice from the conclusions in the second draft PA by also advising 

that a level of 70 ppb would provide little margin of safety for protection of public health, 

particularly for sensitive subpopulations (Frey, 2014, p. 8). In particular, CASAC stated that: 

At 70 ppb, there is substantial scientific certainty of a variety of adverse effects, 

including decrease in lung function, increase in respiratory symptoms, and 

increase in airway inflammation. Although a level of 70 ppb is more protective of 

public health than the current standard, it may not meet the statutory requirement 

to protect public health with an adequate margin of safety (Frey, 2014, p.8). 

However, the committee also acknowledged that “the choice of a level within the range 

recommended based on scientific evidence [i.e., 70 to 60 ppb] is a policy judgment under the 

statutory mandate of the Clean Air Act” (Frey, 2014). 

In summary, compared to the current standard, a revised O3 standard with a level of 70 

ppb would be expected to (1) reduce the occurrence of exposures of concern to O3 concentrations 

that result in respiratory effects in healthy adults (at or above 60 and 70 ppb) by about 45 to 

95%, almost eliminating the occurrence of multiple exposures at or above 70 ppb; (2) reduce the 

occurrence of moderate-to-large O3-induced lung function decrements (FEV1 decrements > 10, 

15, 20%) by about 15 to 35%, most effectively limiting the occurrence of multiple decrements 

and decrements > 15, 20%; (3) more effectively maintain short- and long-term O3 concentrations 

below those present in the epidemiologic studies that reported significant O3 health effect 

associations in locations likely to have met the current standard;47 and (4) reduce the risk of O3-

associated mortality and morbidity, particularly the risk associated with the upper portions of the 

distributions of ambient O3 concentrations. 

 

47 Though epidemiologic studies also provide evidence for O3 health effect associations in locations likely to have 

met a standard with a level of 70 ppb, as discussed below for lower standard levels. 
4-64 

O3 standard level of 65 ppb 

Next, we consider a standard with a level of 65 ppb. A level of 65 ppb is well below 80 

ppb, an O3 exposure concentration that has been reported to elicit a range of respiratory effects 

that includes airway hyperresponsiveness and decreased lung host defense, in addition to lung 

function decrements, airway inflammation, and respiratory symptoms. A standard level of 65 

ppb is also below the lowest exposure concentration at which the combined occurrence of 

respiratory symptoms and lung function decrements has been reported (i.e., 72 ppb), a 

combination judged adverse by the ATS (section 3.1.3). A level of 65 ppb is above the lowest 

exposure concentration demonstrated to result in lung function decrements large enough to be 

judged an abnormal response by ATS, where statistically significant changes in group mean 

responses would be judged to be adverse by ATS, and which the CASAC has indicted could be 

adverse in people with lung disease (i.e., 60 ppb). A level of 65 ppb is also above the lowest 

exposure concentration at which pulmonary inflammation has been reported in healthy adults 

(i.e., 60 ppb). 

Compared to the current standard and a revised standard with a level of 70 ppb, the 

HREA estimates that a standard with a level of 65 ppb would reduce exposures of concern to the 

range of O3 benchmark concentrations analyzed (i.e., 60, 70, and 80 ppb). The HREA estimates 

that meeting a standard with a level of 65 ppb would eliminate exposures of concern at or above 

80 ppb in the urban case study areas. Such a standard is estimated to allow far less than 1% of 

children in the urban case study areas to experience one or more exposures of concern at or 

above the 70 ppb benchmark level, even in the worst-case years and locations, and is estimated 

to eliminate the occurrence of two or more exposures at or above 70 ppb (Table 4-4). 

In addition, on average over the years 2006 to 2010, a standard with a level of 65 ppb is 

estimated to allow between 0 and about 4% of children (including asthmatic children) in urban 

case study areas to experience exposures of concern at or above 60 ppb, which CASAC has 

indicated is an appropriate exposure of concern for people with asthma, including children. This 

reflects an 80% reduction (on average across areas), relative to the current standard. A standard 

with a level of 65 ppb is estimated to allow less than 1% of children to experience two or more 

exposures of concern at or above 60 ppb (> 90% reduction, compared to current standard). In the 

worst-case location and year, about 10% of children are estimated to experience one or more 

exposures of concern at or above 60 ppb, with about 3% estimated to experience two or more 

such exposures (Table 4-4). 

Compared to the current standard and a revised standard with a level of 70 ppb, the 

HREA estimates that a standard with a level of 65 ppb would also reduce the occurrence of O3-

induced lung function decrements. A level of 65 ppb is estimated to allow about 4% or less of 

children to experience moderate O3-induced FEV1 decrements ≥ 15% (50% reduction, compared 
4-65 

to current standard), even considering the worst-case location and year. Such a standard is 

estimated to allow about 2% or less of children to experience two or more such decrements. A 

standard set at 65 ppb is estimated to allow about 1% or less of children to experience large O3-

induced lung function decrements (i.e., FEV1 decrement ≥ 20%), even in the worst-case year and 

location (Table 4-5). 

On average over the years 2006 to 2010, a standard with a level of 65 ppb is estimated to 

allow about 3 to 15% of children to experience one or more moderate O3-induced lung function 

decrements (i.e., FEV1 decrement ≥ 10%), which CASAC has indicated could be adverse for 

people with lung disease. This reflects an average reduction of about 30%, relative to the current 

standard. A standard with a level of 65 ppb is also estimated to allow about 1 to 9% of children 

in the urban case study areas to experience two or more such decrements (37% reduction, 

compared to current standard). In the worst-case location and year, a standard set at 65 ppb is 

estimated to allow up to about 18% of these children to experience one or more moderate O3-

induced lung function decrements ≥ 10%, and up to 11% to experience two or more such 

decrements (Table 4-5). 

With regard to O3 epidemiologic studies we note that, compared to a standard with a level 

of 70 ppb, a revised standard with a level of 65 ppb would more effectively maintain short-term 

O3 concentrations below those present in the epidemiologic studies that reported significant O3

health effect associations in locations likely to have met the current standard. In particular, 

Katsouyanni et al. (2009) reported statistically significant associations with mortality based on 

air quality in 12 Canadian cities, most of which would likely have met a standard with a level of 

70 ppb over the entire study period but violated a revised standard with a level of 65 ppb or 

below over at least part of that period (Table 4-1). This analysis suggests that while the current 

standard or a standard with a level of 70 ppb would allow the ambient O3 concentrations in most 

of the study locations that provided the basis for the association with mortality in this study, a 

revised O3 standard with a level at or below 65 ppb would require reductions in those ambient O3

concentrations. As discussed above for a level of 70 ppb, this analysis does not provide 

information on the extent to which O3-associated mortality would persist upon meeting an O3

standard with a level of 65 ppb, or on the extent to which standard levels below 65 ppb could 

further reduce the incidence of this mortality.48

With regard to long-term O3 concentrations, as for 70 ppb (above) we evaluate the “longterm” O3 metrics reported to be associated with mortality or morbidity in recent epidemiologic 

 

48 For the other multicity studies identified in Table 4-1 (Cakmak et al., 2006; Stieb et al., 2009; Katsouyanni et al., 

2009 (for hospital admissions)), and for the study by Bell et al. (2006) (for the 30 ppb cut point) (Table 4-2), the 

majority of study locations would have met a standard with a level of 65 ppb. 
4-66 

studies (e.g., seasonal averages of 1-hour or 8-hour daily max concentrations). Compared to the 

current standard or a revised O3 standard with a level of 70 ppb, a revised standard with a level 

of 65 ppb would be expected to further reduce the risk of respiratory mortality associated with 

long-term O3 concentrations, based on information from the study by Jerrett et al. (2009).49 In 

addition, a standard with a level of 65 ppb would be expected to more effectively maintain longterm O3 concentrations below those where the study by Jerrett et al. (2009) indicates the most 

confidence in the reported association with respiratory mortality. Specifically, air quality 

analyses indicate this to be the case in 10 out of the 12 urban case study areas for a level of 65 

ppb, compared to 6 out of 12 areas for the current standard and 9 out of 12 for a standard with a 

level of 70 ppb (Table 4-3). Finally, a revised standard with a level of 65 ppb would be expected 

to further reduce long-term O3 concentrations based on the types of metrics that have been 

reported in recent epidemiologic studies to be associated with respiratory morbidity (i.e., 

seasonal averages of daily maximum 8-hour concentrations). 

In further considering the potential implications of epidemiology studies for alternative 

standard levels, we note estimates of total mortality associated with short-term O3.

50 As 

discussed above, we consider estimates of total risk (i.e., based on the full distributions of 

ambient O3 concentrations) and estimates of risk associated with O3 concentrations in the upper 

portions of ambient distributions. With regard to total risk we note that, when summed across 

urban case study areas, a standard with a level of 65 ppb is estimated to reduce the number of 

deaths associated with short-term O3 exposures by about 13% (2007) and 9% (2009), compared 

to the current standard.51 For area-wide concentrations at or above 40 ppb, a standard level of 65 

ppb is estimated to reduce O3-associated deaths by almost 50% compared to the current standard, 

when summed across cities. For area-wide concentrations at or above 60 ppb, a standard level of 

65 ppb is estimated to reduce O3-associated deaths by more than 80% (Figure 4-13). 

As discussed above, although CASAC concluded that the scientific evidence supports 

considering standard levels as high as 70 ppb, it also concluded that a level of 70 ppb would 

provide little margin of safety (Frey, 2014, p. 8). In support of its policy advice that the level 

should be set below 70 ppb, CASAC noted that an alternative standard with a level of 65 ppb 

would further reduce, though not eliminate, the frequency of lung function decrements > 15% 

 

49 Though as discussed above, we note the lower confidence we place in these risk results (U.S. EPA, 2014a, section 

9.6). 

50 As discussed above, compared to the weight given to the evidence and to HREA estimates of exposures of 

concern and lung function risks, we place relatively less weight on epidemiologic-based risk estimates. 51 A standard with a level of 65 ppb is also estimated to reduce respiratory mortality associated with long-term O3

concentrations in urban case study areas. However, given uncertainties associated with these risk estimates, as 

discussed above, we give them limited weight. 
4-67 

and would lead to lower frequency of short-term premature mortality (i.e., compared to a 

standard with a level of 70 ppb) (Frey, 2014, p. 8). 

In summary, compared to a standard with a level of 70 ppb, a revised standard with a 

level of 65 ppb would be expected to (1) further reduce the occurrence of exposures of concern 

(by about 80 to 100% compared to the current standard), decreasing exposures at or above 60 

ppb and almost eliminating exposures at or above 70 and 80 ppb; (2) further reduce the 

occurrence of FEV1 decrements > 10, 15, and 20% (by about 30 to 65%, compared to the current 

standard); (3) more effectively maintain short- and long-term O3 concentrations below those 

present in the epidemiologic studies that reported significant O3 health effect associations in 

locations likely to have met the current standard;52 and (4) further reduce the risk of O3-

associated mortality and morbidity, particularly the risk associated with the upper portion of the 

distribution of ambient O3 concentrations. 

O3 standard level of 60 ppb 

We next consider a standard with a level of 60 ppb. A level of 60 ppb is well below the 

O3 exposure concentration that has been reported to elicit a wide range of potentially adverse 

respiratory effects in healthy adults (i.e., 80 ppb). A level of 60 ppb is also below the 

concentration where the combined occurrence of respiratory symptoms and lung function 

decrements was observed, a combination judged adverse by the ATS (i.e., 72 ppb, discussed in 

section 3.1.3). A level of 60 ppb corresponds to the lowest exposure concentration demonstrated 

to result in lung function decrements that are large enough to be judged an abnormal response by 

ATS, that meet ATS criteria for adversity based on a downward shift in the distribution of FEV1, 

and that the CASAC indicated could be adverse in people with lung disease. A level of 60 ppb 

also corresponds to the lowest exposure concentration at which pulmonary inflammation has 

been reported in controlled human exposure studies. 

Based on the HREA analyses of O3 exposures of concern, a standard with a level of 60 

ppb is estimated to eliminate exposures of concern at or above the 70 and 80 ppb benchmark 

concentrations and to be more effective than the higher standard levels at limiting exposures of 

concern at or above 60 ppb. On average over the years 2006 to 2010, a standard with a level of 

60 ppb is estimated to allow between 0 and about 1% of children, including asthmatic children, 

in urban case study areas to experience exposures of concern at or above 60 ppb, which CASAC 

indicated is an appropriate exposure of concern for asthmatic children. This reflects a 96% 

reduction (on average across areas), compared to the current standard. A standard with a level of 

 

52Though epidemiologic studies also provide evidence for O3 health effect associations in locations likely to have 

met a standard with a level of 65 ppb, as discussed below for a level of 60 ppb. 
4-68 

60 ppb is estimated to allow virtually no children to experience two or more exposures of 

concern at or above 60 ppb. In the worst-case location and year, about 2% of children are 

estimated to experience exposures of concern at or above 60 ppb, with far less than 1% estimated 

to experience two or more such exposures (Table 4-4). 

Based on the HREA analyses of O3-induced lung function decrements, a standard with a 

level of 60 ppb would be expected to be more effective than a level of 70 or 65 ppb at limiting 

the occurrence of O3-induced lung function decrements. A standard with a level of 60 ppb is 

estimated to allow about 2% or less of children in the urban case study areas to experience one or 

more moderate O3-induced FEV1 decrements ≥ 15% (almost 70% reduction, compared to current 

standard), and about 1% or less to experience two or more such decrements (3% in the location 

and year with the largest estimates). A standard set at 60 ppb is estimated to allow about 1% or 

less of children to experience large O3-induced lung function decrements (i.e., FEV1 decrement ≥

20%), even in the worst-case locations and year (Table 4-5). 

On average over the years 2006 to 2010, a standard with a level of 60 ppb is estimated to 

allow about 5 to 11% of children in the urban case study areas to experience one or more 

moderate O3-induced lung function decrements that CASAC indicated could be adverse for 

people with lung disease (i.e., FEV1 decrements ≥ 10%). This reflects an average reduction of 

about 45%, compared to current standard. A standard with a level of 60 ppb is also estimated to 

allow about 2 to 6% of children in these areas to experience two or more such decrements (51% 

reduction, compared to current standard). In the worst-case location and year, a standard set at 60 

ppb is estimated to allow up to about 13% of children to experience one or more moderate O3-

induced FEV1 decrements ≥ 10%, and 7% to experience two or more such decrements (Table 4-

5). 

With regard to O3 epidemiologic studies we note that, compared to a standard with a level 

of 70 or 65 ppb, a revised standard with a level of 60 ppb would more effectively maintain shortterm O3 concentrations below those present in the epidemiologic studies that reported significant 

O3 health effect associations in locations likely to have met the current standard. Specifically, in 

all of the U.S. and Canadian epidemiologic studies evaluated, the majority of study cities had 

ambient O3 concentrations that would likely have violated a standard with a level of 60 ppb. 

Thus, none of the U.S. and Canadian epidemiologic studies analyzed provide evidence for O3

health effect associations when the majority of study locations would likely have met a standard 

with a level of 60 ppb (Tables 4-1 and 4-2). As discussed above, while this analysis does not 

provide information on the extent to which the O3-associated morbidity or mortality would 

persist upon meeting an O3 standard with a level of 60 ppb, it suggests that a revised O3 standard 

with a level of 60 ppb would require reductions in the ambient O3 concentrations that provided 

the basis for those health effect associations. 
4-69 

With regard to long-term O3 concentrations, compared to the current standard or a revised 

O3 standard with a level of 70 or 65 ppb, a revised standard with a level of 60 ppb would be 

expected to further reduce the risk of respiratory mortality associated with long-term O3

concentrations, based on information from the study by Jerrett et al. (2009).53 In addition, a 

standard with a level of 60 ppb would be expected to more effectively maintain long-term O3

concentrations below those where the study by Jerrett et al. (2009) indicates the most confidence 

in the reported association with respiratory mortality. Specifically, air quality analyses indicate 

this to be the case in all of the urban case study areas evaluated at a level of 60 ppb, compared to 

6 out of 12 areas for the current standard, 9 out of 12 for a standard with a level of 70 ppb, and 

10 out of 12 for a standard with a level of 65 ppb (Table 4-3). Finally, a revised standard with a 

level of 60 ppb would be expected to further reduce long-term O3 concentrations based on the 

types of metrics that have been reported in recent epidemiologic studies to be associated with 

respiratory morbidity (i.e., seasonal averages of daily maximum 8-hour concentrations). 

In further considering the potential implications of epidemiology studies for alternative 

standard levels, we note estimates of total mortality associated with short-term O3

concentrations.54 As discussed above, we consider estimates of total risk (i.e., based on the full 

distributions of ambient O3 concentrations) and estimates of risk associated with O3

concentrations in the upper portions of ambient distributions. With regard to total risk we note 

that, when summed across urban case study areas, a standard with a level of 60 ppb is estimated 

to reduce the number of deaths associated with short-term O3 exposures by about 15% (2007) 

and 11% (2009), compared to the current standard (Figure 4-13).55 For area-wide concentrations 

at or above 40 ppb, a standard with a level set at 60 ppb is estimated to reduce O3-associated 

deaths by almost 60% compared to the current standard. For area-wide concentrations at or 

above 60 ppb, a standard level of 60 ppb is estimated to reduce O3-associated deaths by over 

95% compared to the current standard (Figure 4-13). 

Relative to the current standard, or alternative O3 standards with levels of 70 or 65 ppb, 

CASAC stated the following: 

The frequency of lung function decrements and premature mortality from shortterm exposure to ozone decreases even further when the alternative standard is 

lowered to 60 ppb (Frey, 2014, p.8). 

 

53 Though as discussed above, we note the lower confidence we place in these risk results (U.S. EPA, 2014a, section 

9.6). 

54 As discussed above, compared to the weight given to the evidence and to HREA estimates of exposures of 

concern and lung function risks, we place relatively less weight on epidemiologic-based risk estimates. 55 A standard with a level of 60 ppb is also estimated to reduce respiratory mortality associated with long-term O3

concentrations in urban case study areas. However, given uncertainties associated with these risk estimates, as 

discussed above, we give them limited weight. 
4-70 

CASAC also concluded that “the recommended lower bound of 60 ppb would certainly offer 

more public health protection than levels of 70 ppb or 65 ppb and would provide an adequate 

margin of safety” (Frey, 2014, p. ii). 

In summary, compared to a standard with a level of 70 or 65 ppb, a revised standard with 

a level of 60 ppb would be expected to (1) further reduce the occurrence of exposures of concern 

(by about 95 to 100% compared to the current standard), almost eliminating exposures at or 

above 60 ppb; (2) further reduce the occurrence of FEV1 decrements > 10, 15, and 20%, (by 

about 45 to 85% compared to the current standard); (3) more effectively maintain short- and 

long-term O3 concentrations below those present in the epidemiologic studies that reported 

significant O3 health effect associations in locations likely to have met the current standard;56 and 

(4) further reduce the risk of O3-associated mortality and morbidity, particularly the risk 

associated with the upper portion of the distribution of ambient O3 concentrations. 

4.7 KEY UNCERTAINTIES AND AREAS FOR FUTURE RESEARCH AND 

DATA COLLECTION 

It is important to highlight the uncertainties associated with establishing standards for O3

during and after completion of the NAAQS review process. Research needs go beyond what is 

necessary to understand health effects, population exposures, and risks of exposure for purposes 

of setting standards. Research can also support the development of more efficient and effective 

control strategies. In this section, we highlight areas for future health-related research, model 

development, and data collection activities to address these uncertainties and limitations in the 

current body of scientific evidence. 

As has been presented and discussed in the ISA, particularly chapters 4 through 7, the 

scientific body of evidence informing our understanding of health effects associated with longand short-term exposures to O3 has been broadened and strengthened since the O3 NAAQS 

review completed in 2008. Still, we have concluded that O3 health research needs and priorities 

have not changed substantially since the 2007 O3 Staff Paper (EPA 2007). Key uncertainties and 

research needs that continue to be high priority for future reviews of the health-based standards 

are identified below: 

(1) An important aspect of risk characterization and decision making for air quality 

standard levels for the O3 NAAQS is the characterization of the shape of exposure-response 

functions for O3, including the identification of potential population threshold levels. Recent 

controlled human exposure studies of measurable lung function effects provide evidence for a 

smooth dose-response curve without evidence of a threshold for exposures between 40 and 120 

 

56As discussed above, these studies do not provide information on the extent to which O3 health effect associations 

would persist following reductions in ambient O3 concentrations in order to meet a standard with a level of 60 ppb. 
4-71 

ppb O3 (US EPA, 2013, Figure 6-1). Considering the importance of estimating health risks in the 

range below 80 ppb O3, additional research is needed to evaluate responses in healthy and 

especially people with asthma in the range of 40 to 70 ppb for 6-8 hour exposures while engaged 

in moderate exertion. 

(2) Similarly, for health endpoints reported in epidemiologic studies such as hospital 

admissions, ED visits, and premature mortality, an important aspect of characterizing risk is the 

shape of concentration-response functions for O3, including identification of potential population 

threshold levels. Most of the recent studies and analyses continue to show no evidence for a clear 

threshold in the relationships between O3 concentrations commonly observed in the U.S. during 

the O3 season and these health endpoints, though evidence indicates less certainty in the shape of 

the concentration-response curve at the lower end of the distribution of O3 concentrations. 

However, there continues to be heterogeneity in the O3-mortality relationship across cities (or 

regions), including effect modifiers that are also expected to vary regionally, which are sources 

of uncertainty. Additionally, whether or not exposure errors, misclassification of exposure, or 

potential impacts of other copollutants may be obscuring potential population thresholds is still 

unknown. 

(3) The extent to which the broad mix of photochemical oxidants and more generally 

other copollutants in the ambient air (e.g., PM, NO2, SO2, etc.) may play a role in modifying or 

contributing to the observed associations between ambient O3 and various morbidity effects and 

mortality continues to be an important research question. Ozone has long been known as an 

indicator of health effects of the entire photochemical oxidant mix in the ambient air and has 

served as a surrogate for control purposes. A better understanding of sources of the broader 

pollutant mix, of human exposures, and of how other pollutants may modify or contribute to the 

health effects of O3 in the ambient air, and vice versa, is needed to better inform future NAAQS 

reviews. 

(4) As epidemiologic research has continued to be an important factor in assessing the 

public health impacts of O3, methodological issues in epidemiologic studies have received 

greater visibility and scrutiny. There remains a need to further examine alternative modeling 

specifications and control of time-varying factors, and to better understand the role of 

copollutants in the ambient air. Additionally, there remains uncertainty around the role of 

temperature as a potential confounder or effect modifier in epidemiologic models. 

(5) Recent animal toxicological evidence, combined with limited evidence from 

controlled human exposure studies of cardiovascular morbidity and epidemiologic studies of 

cardiovascular mortality, have provided evidence of both direct and indirect effects on the 

cardiovascular system. However, additional work will need to examine biologically plausible 

mechanisms of cardiovascular effects, expand upon preliminary evidence from controlled human 
4-72 

exposure studies, address inconsistencies observed in epidemiologic studies of cardiovascular 

morbidity, and determine the extent to which O3 is directly implicated or works together with 

other pollutants in causing adverse cardiovascular effects in both at-risk and the general 

populations. 

(6) Most epidemiologic studies of short-term exposure effects have employed time-series 

or case-crossover study designs and have been conducted in large populations. These study 

designs remain subject to uncertainty due to use of ambient fixed-site data serving as a surrogate 

for ambient exposures, and to the difficulty of determining the impact of any single pollutant 

among the mix of pollutants in the ambient air. Measurements made at stationary outdoor 

monitors have been used as independent variables for air pollution, but the accuracy with which 

these measurements actually reflect subjects’ exposure is not yet fully understood. Also, 

additional research is needed to improve the characterization of the degree to which discrepancy 

between stationary monitor measurements and actual pollutant exposures introduces error into 

statistical estimates of pollutant effects in epidemiologic studies. 

(7) Recent studies of “long-term” O3 often evaluate associations with daily maximum 

concentrations, averaged over the O3 season. Research is needed to better understand the extent 

to which health effects associated with such long-term metrics are attributable to long-term 

average concentrations versus the repeated occurrence of daily maximum concentrations. 

(8) Improved understanding of human exposures to ambient O3 and to related 

copollutants is an important research need. Population-based information on human exposure for 

healthy adults and children and at-risk populations, including people with asthma, to ambient O3

concentrations, including exposure information in various microenvironments, is needed to better 

evaluate current and future O3 exposure models. Such information is needed for sufficient 

periods to facilitate evaluation of exposure models throughout the O3 season. 

(9) Information is needed to improve inputs to current and future population-based O3

exposure and health risk assessment models. Collection of time-activity data over longer time 

periods is needed to reduce uncertainty in the modeled exposure distributions that form an 

important part of the basis for decisions regarding NAAQS for O3 and other air pollutants. 

Research addressing energy expenditure and associated breathing rates in various population 

groups, particularly healthy children and children with asthma, in various locations, across the 

spectrum of physical activity, including sleep to vigorous exertion, is needed. 

(10) An important consideration in the O3 NAAQS review is the characterization of 

background levels. There still remain substantial uncertainties in the characterization of 8-hour 

daily max O3 background concentrations. Further research to improve the evaluation of the 

global and regional models which have been used to characterize estimates of background levels 
4-73 

would improve understanding of the role of non-U.S. anthropogenic emissions on O3 levels over 

the U.S. 

4.8 SUMMARY OF STAFF CONCLUSIONS ON PRIMARY STANDARD 

In this section, we summarize our conclusions regarding the primary O3 standard. These 

conclusions are informed by our consideration of the available scientific evidence as assessed in 

the ISA, air quality/exposure/risk information assessed in the final HREA, recommendations and 

advice received from CASAC, and comments received from members of the public. 

As an initial matter in this PA, staff concludes that reducing ambient O3 concentrations to 

meet the current standard will provide important improvements in public health protection. This 

initial conclusion is based on (1) the strong body of scientific evidence indicating a wide range of 

adverse health outcomes attributable to exposures to O3 concentrations found in the ambient air 

and (2) estimates indicating decreased O3 exposures and health risks upon meeting the current 

standard, compared to recent air quality. Strong support for this conclusion is provided by the 

available health evidence, and by HREA estimates of exposures to O3 concentrations shown to 

result in respiratory effects in healthy adults (exposures of concern > 60, 70 and 80 ppb); O3-

induced lung function risks (FEV1 decrements ≥ 10, 15 and 20%); and O3-associated mortality 

and morbidity risks. 

Staff further concludes that the O3-attributable health effects estimated to be allowed by 

air quality that meets the current primary standard can reasonably be judged important from a 

public health perspective. This conclusion is based on consideration of the scientific evidence 

assessed in the ISA, including controlled human exposure studies reporting abnormal or adverse 

respiratory effects following exposures to O3 concentrations below the level of the current 

standard and epidemiologic studies indicating associations with morbidity and mortality for air 

quality that would meet the current standard. This conclusion is also based on the HREA 

estimates of exposures of concern, lung function risks, and morbidity and mortality risks; on 

advice received from CASAC in their review of draft versions of the PA; on CASAC advice 

received in previous reviews; and on consideration of public comments. Staff reaches the overall 

conclusion that the available health evidence and exposure/risk information calls into question 

the adequacy of the public health protection provided by the current standard. 

Given this conclusion regarding the adequacy of the current standard, staff also reaches 

conclusions for the Administrator’s consideration regarding the elements of alternative primary 

O3 standards that could be supported by the available evidence and exposure/risk information. In 

reaching conclusions about the range of potential alternative standards appropriate for 

consideration, staff is mindful that the Act requires primary standards that, in the judgment of the 

Administrator, are requisite to protect public health with an adequate margin of safety. The 
4-74 

primary standards are to be neither more nor less stringent than necessary. Thus, the Act does not 

require that primary NAAQS be set at zero-risk levels, but rather at levels that reduce risk 

sufficiently to protect public health with an adequate margin of safety. 

The degree of public health protection provided by any NAAQS results from the 

collective impact of the elements of the standard, including the indicator, averaging time, level, 

and form. Staff’s conclusions on each of these elements are summarized below. 

(1) Indicator: It is appropriate to continue to use O3 as the indicator for a standard that is 

intended to address effects associated with exposure to O3, alone or in combination with 

related photochemical oxidants. Based on the available information, staff concludes that 

there is no basis for considering any alternative indicator at this time. Meeting an O3

standard can be expected to provide some degree of protection against potential health 

effects that may be independently associated with other photochemical oxidants, even 

though such effects are not discernible from currently available studies indexed by O3

alone. Staff notes that control of ambient O3 levels is generally understood to provide the 

best means of controlling photochemical oxidants, and thus of protecting against effects 

that may be associated with individual species and/or the broader mix of photochemical 

oxidants. 

(2) Averaging time: It is appropriate to consider continuing to use an 8-hour averaging time 

for the primary O3 standard. 

(a) Staff concludes that an 8-hour averaging time remains appropriate for addressing 

health effects associated with short-term exposures to ambient O3. An 8-hour 

averaging time is similar to the exposure periods evaluated in controlled human 

exposure studies, including recent studies reporting respiratory effects following 

exposures to O3 concentrations below the level of the current standard. In 

addition, epidemiologic studies provide evidence for health effect associations 

with 8-hour O3 concentrations, as well as with 1-hour and 24-hour concentrations. 

A standard with an 8-hour averaging time (combined with an appropriate standard 

form and level) would also be expected to provide substantial protection against 

health effects attributable to 1- and 24-hour exposures. 

(b) Staff also concludes that a standard with an 8-hour averaging time can provide 

protection against respiratory effects associated with longer term O3 exposures. 

Analyses in the HREA show that as air quality is adjusted to just meet the current 
4-75 

or alternative 8-hour standards, most study areas are estimated to experience 

reductions in respiratory mortality associated with long-term O3 concentrations. 

In addition, analyses in this PA indicate that just meeting an 8-hour standard with 

an appropriate level would be expected to maintain long-term O3 concentrations 

(i.e., seasonal average of 1-hour daily max) below those where a key study 

indicates the most confidence in the concentration-response relationship with 

respiratory mortality. In considering other long-term O3 metrics evaluated in 

recent health studies, analyses in the HREA indicate that the large majority of the 

U.S. population lives in locations where reducing NOX emissions would be 

expected to decrease warm season averages of daily 8-hour ambient O3

concentrations, a long-term metric used in several recent studies reporting 

associations with respiratory morbidity. Taken together, these analyses suggest 

that a standard with an 8-hour averaging time, coupled with the current 4th-highest 

form and an appropriate level, could provide appropriate protection against the 

long-term O3 concentrations reported to be associated with respiratory morbidity 

and mortality. 

(3) Form: For an 8-hour O3 standard with a revised level (as discussed below), it is 

appropriate to consider retaining the current form, defined as the 3-year average of the 

annual 4th-highest daily maximum concentration. Staff notes that this form was selected 

in 1997 and 2008 in recognition of the public health protection provided, when coupled 

with an appropriate averaging time and level, combined with the stability provided for 

implementation programs. The currently available evidence and exposure/risk 

information does not call into question these conclusions from previous reviews. 

(4) Level: The available scientific evidence and exposure/risk information provide strong 

support for considering an O3 standard with a revised level in order to increase public 

health protection. Staff concludes that it is appropriate in this review to consider a revised 

standard level within the range of 70 ppb to 60 ppb, reflecting the judgment that a 

standard set within this range could provide an appropriate degree of public health 

protection and would result in important improvements in protection for at-risk 

populations and lifestages. 

 
4-76 

4.9 REFERENCES 

Adams, WC. (1998). Dose-response effect of varied equivalent minute ventilation rates on pulmonary function 

responses during exposure to ozone. Washington, DC: American Petroleum Institute. 

Adams, W. C. (2006) Comparison of chamber 6.6 hour exposures to 0.04-0.08 ppm ozone via square-wave and 

triangular profiles on pulmonary responses. Inhalation Toxicol. 18: 127-136. 

http://dx.doi.org/10.1080/08958370500306107 

Bell, ML; Peng, RD; Dominici, F. (2006). The exposure-response curve for ozone and risk of mortality and the 

adequacy of current ozone regulations. Environ Health Perspect 114: 532-536. 

Brown, JS; Bateson, TF; McDonnell, WF. (2008). Effects of exposure to 0.06 ppm ozone on FEV1 in humans: A 

secondary analysis of existing data. Environ Health Perspect 116: 1023-1026. 

http://dx.doi.org/10.1289/ehp.11396 

Cakmak, S; Dales, RE; Judek, S. (2006). Respiratory health effects of air pollution gases: Modification by education 

and income. Arch Environ Occup Health 61: 5-10. http://dx.doi.org/10.3200/AEOH.61.1.5-10 

Dales, RE; Cakmak, S; Doiron, MS. (2006). Gaseous air pollutants and hospitalization for respiratory disease in the 

neonatal period. Environ Health Perspect 114: 1751-1754. http://dx.doi.org/10.1289/ehp.9044 

Darrow, LA; Klein, M; Sarnat, JA; Mulholland, JA; Strickland, MJ; Sarnat, SE; Russell, AG; Tolbert, PE. (2011). 

The use of alternative pollutant metrics in time-series studies of ambient air pollution and respiratory 

emergency department visits. J Expo Sci Environ Epidemiol 21: 10-19. 

http://dx.doi.org/10.1038/jes.2009.49 

Frey, C. (2014) CASAC Review of the EPA’s Second Draft Policy Assessment for the Review of the Ozone National 

Ambient Air Quality Standards EPA-CASAC-14-004. June 26, 2014. Available online at: 

http://yosemite.epa.gov/sab/sabproduct.nsf/264cb1227d55e02c85257402007446a4/5EFA320CCAD326E8

85257D030071531C/$File/EPA-CASAC-14-004+unsigned.pdf 

Frey, C.; Samet, J.M. (2012). CASAC Review of the EPA’s Policy Assessment for the Review of the Ozone National 

Ambient Air Quality Standards (First External Review Draft – August 2012). EPA-CASAC-13-003. 

November 26, 2012. Available online at: 

Islam, T; McConnell, R; Gauderman, WJ; Avol, E; Peters, JM; Gilliland, FD. (2008). Ozone, oxidant defense genes 

and risk of asthma during adolescence. Am J Respir Crit Care Med 177: 388-395. 

http://dx.doi.org/10.1164/rccm.200706-863OC 

Jerrett, M; Burnett, RT; Pope, CA, III; Ito, K; Thurston, G; Krewski, D; Shi, Y; Calle, E; Thun, M. (2009). Longterm ozone exposure and mortality. N Engl J Med 360: 1085-1095. 

http://dx.doi.org/10.1056/NEJMoa0803894 

Katsouyanni, K; Samet, JM; Anderson, HR; Atkinson, R; Le Tertre, A; Medina, S; Samoli, E; Touloumi, G; 

Burnett, RT; Krewski, D; Ramsay, T; Dominici, F; Peng, RD; Schwartz, J; Zanobetti, A. (2009). Air 

pollution and health: A European and North American approach (APHENA). (Research Report 142). 

Boston, MA: Health Effects Institute. http://pubs.healtheffects.org/view.php?id=327 

Kim, CS; Alexis, NE; Rappold, AG; Kehrl, H; Hazucha, MJ; Lay, JC; Schmitt, MT; Case, M; Devlin, RB; Peden, 

DB; Diaz-Sanchez, D. (2011). Lung function and inflammatory responses in healthy young adults exposed 

to 0.06 ppm ozone for 6.6 hours. Am J Respir Crit Care Med 183: 1215-1221. 

http://dx.doi.org/10.1164/rccm.201011-1813OC 

Lin, S; Liu, X; Le, LH; Hwang, SA. (2008b). Chronic exposure to ambient ozone and asthma hospital admissions 

among children. Environ Health Perspect 116: 1725-1730. http://dx.doi.org/10.1289/ehp.11184 
4-77 

Mar, TF; Koenig, JQ. (2009). Relationship between visits to emergency departments for asthma and ozone exposure 

in greater Seattle, Washington. Ann Allergy Asthma Immunol 103: 474-479. 

McDonnell, WF; Stewart, PW; Smith, MV. (2010). Prediction of ozone-induced lung function responses in humans. 

Inhal Toxicol 22: 160-168. http://dx.doi.org/10.3109/08958370903089557 

Salam, MT; Islam, T; Gauderman, WJ; Gilliland, FD. (2009). Roles of arginase variants, atopy, and ozone in 

childhood asthma. J Allergy Clin Immunol 123: 596-602. http://dx.doi.org/10.1016/j.jaci.2008.12.020 

Samet, J.M. (2010) Review of EPA’s Proposed Ozone National Ambient Air Quality Standard (Federal Register, 

Vol. 75, Nov. 11, January 19, 2010). EPA-CASAC-10-007. February 19, 2010. Available online at: 

 http://yosemite.epa.gov/sab/sabproduct.nsf/264cb1227d55e02c85257402007446a4/610BB57CFAC8A

 41C852576CF007076BD/$File/EPA-CASAC-10-007-unsigned.pdf 

Samet, J.M. (2011) Clean Air Scientific Advisory Committee (CASAC) Response to Charge Questions on the 

Reconsideration of the 2008 Ozone National Ambient Air Quality Standards. EPA-CASAC-11-004. 

March 30, 2011. Available online at: 

http://yosemite.epa.gov/sab/sabproduct.nsf/0/F08BEB48C1139E2A8525785E006909AC/$File/EPACASAC-11-004-unsigned+.pdf 

Sasser, E. (2014) Memo Responding to Request for Revised Ozone HREA Chapter 7 Appendix Tables. May 9, 

2014. Available at http://www.epa.gov/ttn/naaqs/standards/ozone/s_o3_2008_rea.html 

Schelegle, ES; Morales, CA; Walby, WF; Marion, S; Allen, RP. (2009). 6.6-hour inhalation of ozone concentrations 

from 60 to 87 parts per billion in healthy humans. Am J Respir Crit Care Med 180: 265-272. 

http://dx.doi.org/10.1164/rccm.200809-1484OC 

Smith, RL; Xu, B; Switzer, P. (2009). Reassessing the relationship between ozone and short-term mortality in U.S. 

urban communities. Inhal Toxicol 21: 37-61. http://dx.doi.org/10.1080/08958370903161612 

Stieb, DM; Szyszkowicz, M; Rowe, BH; Leech, JA. (2009). Air pollution and emergency department visits for 

cardiac and respiratory conditions: A multi-city time-series analysis. Environ Health Global Access Sci 

Source 8: 25. http://dx.doi.org/10.1186/1476-069X-8-25 

U.S. EPA (2007). Review of the National Ambient Air Quality Standards for Ozone: Policy Assessment of 

Scientific and Technical Information, OAQPS Staff Paper. EPA-452/R-07-007 

U.S. Environmental Protection Agency. (2012). Health Risk and Exposure Assessment for Ozone, First External 

Review Draft, U.S. Environmental Protection Agency, Research Triangle Park, NC. EPA 452/P-12-001. 

U.S. Environmental Protection Agency. (2013). Integrated Science Assessment for Ozone and Related 

Photochemical Oxidants (Final Report). U.S. Environmental Protection Agency, Washington, DC, 

EPA/600/R-10/076F. 

U.S. Environmental Protection Agency. (2014). Health Risk and Exposure Assessment for Ozone. Office of Air 

Quality Planning and Standards, Research Triangle Park, NC. EPA-452/R-14-004a. Available at 

http://www.epa.gov/ttn/naaqs/standards/ozone/s_o3_index.html 
5-1 

5 ADEQUACY OF THE CURRENT SECONDARY STANDARD

This chapter presents staff’s considerations and conclusions regarding the adequacy of 

the current secondary O3 NAAQS. In doing so, we pose the following overarching question: 

 Does the currently available scientific evidence- and exposure/risk-based 

information, as reflected in the ISA and WREA, support or call into question the 

adequacy and appropriateness of the protection afforded by the current 

secondary O3 standard? 

In addressing this overarching question, we pose a series of more specific questions, as 

discussed in sections 5.1 through 5.5 below. We consider the nature of O3-induced effects, 

including the nature of the exposures that drive the biological and ecological response and 

related biologically relevant exposure metrics (section 5.1); the scientific evidence and 

exposure/risk information, including that for associated ecosystem services, regarding (a) tree 

growth, productivity and carbon storage (section 5.2), (b) crop yield loss (section 5.3), (c) visible 

foliar injury (section 5.4), and (d) other welfare effects (section 5.5). Section 5.6 describes 

advice and recommendations received from CASAC. In section 5.7, we revisit the overarching 

question of this chapter and present staff conclusions on the adequacy and appropriateness of the 

current secondary standard. 

5.1 NATURE OF EFFECTS AND BIOLOGICALLY RELEVANT EXPOSURE 

METRIC 

 Does the current evidence alter our conclusions from the previous review 

regarding the nature of O3-induced welfare effects? 

As discussed further below, the current body of O3 welfare effects evidence confirms and 

strengthens the conclusions reached in the last review on the nature of O3-induced welfare 

effects. Ozone’s phytotoxic effects were first identified on grape leaves in a study published in 

1958 (Richards et al., 1958). In the more than fifty years that have followed, extensive research 

has been conducted both in and outside of the U.S. to examine the impacts of O3 on plants and 

their associated ecosystems, since “of the phytotoxic compounds commonly found in the 

ambient air, O3 is the most prevalent, impairing crop production and injuring native vegetation 

and ecosystems more than any other air pollutant” (U.S. EPA, 1989, 1996). Recent studies, 

assessed in the ISA, together with this longstanding and well established vegetation effects 

literature, further contribute to the coherence and consistency of the vegetation effects evidence. 

In assessing the strength of the evidence, it is important to note that different types of 

studies can provide different types of information, each with different associated uncertainties 
5-2 

(U.S. EPA, 2013, Chapter 9, section 9.2). Controlled chamber studies are the best method for 

isolating or characterizing the role of O3 in inducing the observed plant effects, and in assessing 

plant response to O3 at the finer scales (U.S. EPA, 2013, Chapter 9, section 9.3). Recent 

controlled studies have focused on a variety of plant responses to O3 including: 1) the underlying 

mechanisms as they relate to growth, productivity and carbon storage including: reduced carbon 

dioxide uptake due to stomatal closure (U.S. EPA 2013, section 9.3.2.1); 2) the upregulation of 

genes associated with plant defense, signaling, hormone synthesis and secondary metabolism 

(U.S. EPA 2013, section 9.3.3.2); 3) the down regulation of genes related to photosynthesis and 

general metabolism (U.S. EPA 2013, section 9.3.3.2); 4) the loss of carbon assimilation capacity 

due to declines in the quantity and activity of key proteins and enzymes (U.S. EPA, 2013, section 

9.3.5.1); and 5) the negative impacts on the efficiency of the photosynthetic light reactions (U.S. 

EPA, 2013, section 9.3.5.1). As described in the ISA, these new studies “have increased 

knowledge of the molecular, biochemical and cellular mechanisms occurring in plants in 

response to O3”, adding “to the understanding of the basic biology of how plants are affected by 

oxidative stress...” (U.S. EPA, 2013, p. 9-11). The ISA further concluded that controlled studies 

“have clearly shown that exposure to O3 is causally linked to visible foliar injury, decreased 

photosynthesis, changes in reproduction, and decreased growth” in many species of vegetation 

(U.S. EPA 2013, p. 1-15). 

Such effects at the plant scale can also be linked to an array of effects at larger spatial 

scales. For example, recent field studies at larger spatial scales, together with previously 

available evidence, support the controlled exposure study results and indicate that “ambient O3

exposures can affect ecosystem productivity, crop yield, water cycling, and ecosystem 

community composition” (U.S. EPA, 2013, p. 1-15; Chapter 9, section 9.4). 

The ISA summarizes the coherence across the full range of effects, from the least serious 

to the most serious, as follows (U.S. EPA, 2013, p. 1-8): 

The welfare effects of O3 can be observed across spatial scales, starting at the 

subcellular and cellular level, then the whole plant and finally, ecosystem-level 

processes. Ozone effects at small spatial scales, such as the leaf of an individual 

plant, can result in effects along a continuum of larger spatial scales. These 

effects include altered rates of leaf gas exchange, growth, and reproduction at the 

individual plant level, and can result in broad changes in ecosystems, such as 

productivity, carbon storage, water cycling, nutrient cycling, and community 

composition. 

Based on its assessment of this extensive body of science, the ISA determined that, with 

respect to vegetation and ecosystems, a causal relationship exists between exposure to O3 in 

ambient air and visible foliar injury effects on vegetation, reduced vegetation growth, reduced 
5-3 

productivity in terrestrial ecosystems, reduced yield and quality of agricultural crops and 

alteration of below-ground biogeochemical cycles (U.S. EPA 2013, Table 1-2). Additionally, the 

ISA determined that a likely to be causal relationship exists between exposures to O3 in ambient 

air and reduced carbon sequestration in terrestrial ecosystems, alteration of terrestrial ecosystem 

water cycling and alteration of terrestrial community composition (U.S. EPA, 2013, Table 1-2). 

With regard to the relationship between O3 and radiative forcing and climate change, the ISA 

determined that there is a causal relationship between changes in tropospheric O3 concentrations 

and radiative forcing, and likely to be a causal relationship between changes in tropospheric O3

concentrations and effects on climate (U.S. EPA, 2013, p. 1-13, and Table 1-3). From this set of 

effects that the ISA has concluded to be causally or likely causally related to O3 in ambient air, 

we focus the discussion in the PA primarily on: 1) impacts on tree growth, productivity and 

carbon storage; 2) crop yield loss; 3) visible foliar injury. Each of these discussions also 

includes where appropriate, a discussion of any known or anticipated impacts that such 

individual plant or species level effects could have at larger scales, including ecosystems, and on 

associated ecosystem services. 

In considering the available vegetation effects evidence, we make note of several 

important contextual features that frame our understanding of the science and how it informs our 

evaluation of the adequacy of the protection afforded by the current secondary NAAQS. First, 

we acknowledge that under natural conditions, a variety of factors can either mitigate or 

exacerbate the predicted O3-plant interactions and are recognized sources of uncertainty and 

variability. These include: 1) multiple genetically influenced determinants of O3 sensitivity; 2) 

changing sensitivity to O3 across vegetative growth stages; 3) co-occurring stressors and/or 

modifying environmental factors (U.S. EPA, 2013, section 9.4.8). 

Second, we acknowledge that the species that have been studied for O3 sensitivity 

represent only a fraction of the tens of thousands of plant species that grow in the U.S. (USDA 

NRCS, 2014)1, and that these species were typically selected because of their commercial 

importance (e.g., commodity crop or timber species) or because of observed O3-induced visible 

foliar injury in the field. Of the species known to be sensitive to O3 for foliar injury, 66 species 

have been identified on National Park Service (NPS) and U.S. Fish and Wildlife Service lands2

and a subset of these are used in the USFS biomonitoring program (discussed in section 5.4 

below). A number of these species have also been identified as important to tribal cultural 

practices (see Appendix 5-A). Appendix 7J of the 2007 Staff Paper showed that no state in the 

 

1

 USDA NRCS. 2014. The PLANTS Database (http://plants.usda.gov, 3 January 2014). National Plant Data 

Team, Greensboro, NC 27401-4901 USA. 

2

 See http://www2.nature.nps.gov/air/Pubs/pdf/flag/NPSozonesensppFLAG06.pdf 
5-4 

lower 48 states had less than seven known O3-sensitive plant species, with the majority of states 

having between 11 and 30 (see Appendix 7J–2 in U.S. EPA, 2007). We would not expect this 

information to have changed since the previous review because there has been very little change 

in the list of sensitive species and the occurrence of any of these plant species within a state 

would not be expected to change. With respect to agricultural species, a number of important 

commodity crops such as soybean and additional fruit and vegetable species such as lettuce have 

been shown to be sensitive to O3 for either foliar injury or yield loss (U.S. EPA, 2013, section 

9.4.4.1; Abt Associates, Inc., 1995).

Third, we acknowledge that out of the group of species known to be sensitive to O3, we 

have chosen to focus primarily on species for which we have robust exposure-response (E-R) 

functions for biomass loss and yield loss using the W126 form (i.e., 11 tree and 10 crop species) 

in order to be able to quantitatively relate predicted changes in O3 to predicted changes in plant 

exposures, responses and associated risks.3 However, while we recognize that this small group 

represents only a fraction of all species known or anticipated to be sensitive to O3 in the U.S., we 

also note, as did CASAC, that among the studied species, there is a fairly large range of O3

sensitivities represented, so that it could be reasonable to assume that other non-studied species 

might have sensitivities that fall within or near this range. Specifically, CASAC states “[i]t 

should not be assumed that species of unknown sensitivity are tolerant to ozone. It is more 

appropriate to assume that the sensitivity of species without E-R functions might be similar to 

the range of sensitivity for those species with E-R functions” (Frey, 2014, p. 11). 

Fourth, we acknowledge that in addition to the well-studied effects of biomass loss in 

trees and crops and visible foliar injury in bioindicator plants that we can quantify, numerous 

other more subtle and less easily observed effects occur along the continuum of spatial scales 

that lead to ecosystem effects. While these effects are more difficult to quantify, we 

acknowledge that any secondary standard set to protect the public welfare against the known and 

quantifiable adverse effects to vegetation should also consider the anticipated, but currently 

unquantifiable, potential adverse effects on vegetation, ecosystems and associated services. 

Finally, we further acknowledge that in light of the above, when considering the available 

evidence, we seek to find the right balance between placing weight on the associated 

uncertainties and limitations of the evidence and placing weight on its well-established strength, 

coherence and consistency. In so doing, we note that CASAC, in commenting on section 6.7 

which describes key uncertainties and future research areas, states that “[w]hile these scientific 

research priorities will enhance future scientific reviews of the ozone primary and secondary 

 

3

 There is an E-R function available for a 12th tree species (cottonwood), but this E-R function is considered 

less robust because it is based on the results of a single gradient study (Gregg et al., 2003). 
5-5 

standards, we also make clear that there is sufficient scientific evidence, and sufficient 

confidence in the available research results, to support the advice we have given above for this 

review cycle of the primary and secondary standards” (Frey, 2014, p. iv). 

 Does the current evidence continue to support a cumulative, seasonal exposure 

index as a biological-relevant and appropriate metric for assessment of the 

evidence and exposure/risk information? 

In this review, the ISA assessment of the full body of currently available evidence stated 

the following regarding biological indices (U.S. EPA, 2013, p. 2-44): 

The main conclusions from the 1996 and 2006 O3 AQCDs [Air Quality Criteria 

Documents] regarding indices based on ambient exposure remain valid. These 

key conclusions can be restated as follows: 

 ozone effects in plants are cumulative; 

 higher O3 concentrations appear to be more important than lower 

concentrations in eliciting a response; 

 plant sensitivity to O3 varies with time of day and plant development 

stage; 

 quantifying exposure with indices that cumulate hourly O3 concentrations 

and preferentially weight the higher concentrations improves the 

explanatory power of exposure/response models for growth and yield, 

over using indices based on mean and peak exposure values. 

 The long-standing body of available evidence upon which these conclusions are based, 

provides a wealth of information on aspects of O3 exposure that are important in influencing 

plant response. Specifically, a variety of “factors with known or suspected bearing on the 

exposure-response relationship, including concentration, time of day, respite time, frequency of 

peak occurrence, plant phenology, predisposition, etc.,” have been identified (U.S. EPA, 2013, 

section 9.5.2). In addition, the importance of the duration of the exposure and the relatively 

greater importance of higher concentrations over lower in determining plant response to O3 have 

been consistently well documented (U.S. EPA, 2013, section 9.5.3). Much of this evidence was 

assessed in the 1996 Criteria Document (CD) (U.S. EPA, 1996), while more recent work 

substantiating this evidence is assessed in the subsequent 2006 CD and 2013 ISA. 

 Understanding of the biological basis for plant response to O3 exposure led to the 

development of a large number of “mathematical approaches for summarizing ambient air 

quality information in biologically meaningful forms for O3 vegetation effects assessment 

purposes ...” (U.S. EPA, 2013, section 9.5.3), including those that cumulate exposures over 

some specified period while weighting higher concentrations more than lower (U.S. EPA, 2013, 

section 9.5.2). As with any summary statistic, these exposure indices retain information on 
5-6 

some, but not all, characteristics of the original observations. As discussed in greater detail in 

section 6.2 below, the 1996 CD contained an extensive review of the published literature on 

different types of exposure-response metrics, including comparisons between metrics, from 

which the 1996 Staff Paper built its assessment of forms appropriate to consider in the context of 

the secondary NAAQS review. The result of these assessments was a decision by the EPA to 

focus on cumulative, concentration-weighted indices, which were recognized as the most 

appropriate biologically based metrics to consider in this context, with attention given primarily 

to two cumulative, concentration weighted index forms: SUM06 and W126. The SUM06 index 

is a threshold-based approach described as the sum of all hourly O3 concentrations greater or 

equal to 0.06 ppm observed during a specified daily and seasonal time window (U.S. EPA, 2013, 

section 9.5.2). The W126 index is a non-threshold approach described as the sigmoidally 

weighted sum of all hourly O3 concentrations observed during a specified daily and seasonal 

time window, where each hourly O3 concentration is given a weight that increases from 0 to 1 

with increasing concentration (Lefohn et al., 1988; Lefohn and Runeckles, 1987; U.S. EPA, 

2013, section 9.5.2). 

In both the 1997 and 2008 reviews, the EPA concluded that the risk to vegetation comes 

primarily from cumulative exposures to O3 over a season or seasons4 and proposed, as one policy 

alternative, a secondary standard set in terms of such a form: SUM06 (61 FR 65716) and W126 

(72 FR 37818) in the 1997 and 2008 reviews, respectively. Although in both reviews the policy 

decision was made to set the secondary standard to be identical to a revised primary standard 

(with an 8-hour averaging time), the Administrator, in both cases, also concluded, consistent 

with CASAC advice, that a cumulative, seasonal index was the most biologically relevant way to 

relate exposure to plant growth response (62 FR 38856, 73 FR 16436). Similarly, in the 2010 

proposed reconsideration of the 2008 decision, the EPA proposed to conclude that O3 exposure 

indices that cumulate differentially weighted hourly concentrations are the best candidates for 

relating exposure to plant growth responses and proposed as the only policy option to set the 

secondary standard in terms of one such form, the W126 (75 FR 2938). This approach of 

establishing a secondary standard that was separate and distinct from the primary standard and in 

particular using a cumulative seasonal exposure index such as W126 received strong support 

from CASAC in both 2008 and 2010 reviews (Henderson, 2006, 2008; Samet, 2010), as it has 

again in this review, as discussed in section 5.6 below. 

An alternative to using ambient exposure durations and concentrations to predict plant 

response has been developed in recent years, primarily in Europe, i.e., flux models. While 

 

4

 In describing the form as “seasonal”, the EPA is referring generally to the growing season of O3-sensitive 

vegetation, not to the seasons of the year (i.e., spring, summer, fall, winter). 
5-7 

“some researchers have claimed that using flux models can be used to better predict vegetation 

responses to O3 than exposure-based approaches...” because flux models estimate the ambient 

O3 concentration that actually enters the leaf (i.e., flux or deposition) (U.S. EPA, 2013, p. 9-114), 

it is important to note that “[f]lux calculations are data intensive and must be carefully 

implemented” (U.S. EPA, 2013, p. 9-114). Further, “[t]his uptake-based approach to quantify 

the vegetation impact of O3 requires inclusion of those factors that control the diurnal and 

seasonal O3 flux to vegetation (e.g., climate patterns, species and/or vegetation-type factors and 

site-specific factors)” (U.S. EPA, 2013, p. 9-114). In addition to these data requirements, each 

species has different amounts of internal detoxification potential that may protect species to 

differing degrees. This balance between O3 flux and detoxification processes has been termed 

the “effective flux”. Accordingly, the “models have to distinguish between stomatal and nonstomatal components of O3 deposition to adequately estimate actual concentration reaching the 

target tissue of a plant to elicit a response” and “ ultimately the ‘effective’ flux” (U.S. EPA, 

2013, pp. 9-114). The lack of detailed species- and site-specific data required for flux modeling 

in the U.S. and the lack of understanding of detoxification processes have continued to make this 

technique less viable for use in vulnerability and risk assessments at the national scale in the U.S. 

(U.S. EPA, 2013, section 9.5.4). 

Therefore, consistent with the ISA conclusions regarding the appropriateness of 

considering cumulative exposure indices that preferentially weight higher concentrations over 

lower for predicting O3 effects of concern based on the long-established conclusions and longstanding supporting evidence described above, and in light of continued CASAC support, we 

continue to focus on the aspects of ambient O3 exposures that have biological relevance and the 

biologically relevant exposure indices or metrics that have been designed in light of this 

consideration, i.e., cumulative concentration-weighted indices. In addition, given the lack of any 

information in the current review to the contrary, we therefore again conclude that the current 

evidence, as in recent reviews, continues to support a cumulative, seasonal exposure index as a 

biologically relevant and appropriate metric for assessment of the evidence and exposure/risk 

information, and in particular, the W126 cumulative, seasonal metric (U.S. EPA, 2013, section 

2.6.6.1, section 9.5.2). Such a metric, as stated above, has an “explanatory power” that is 

improved “over using indices based on mean and peak exposure values” (U.S. EPA, 2013, 

section 2.6.6.1, p. 2-44). Thus, as in the WREA, discussions of the effects evidence and 

exposure/risk results in sections 5.2 through 5.5 of this PA are provided in terms of the W126 

index, where available. 

 What paradigm is being used to consider which of the known or anticipated 

O3-induced effects have the potential to be adverse to the public welfare? 
5-8 

The Clean Air Act (CAA), in section 109 (b) (2) requires that “[a]ny national secondary 

ambient air quality standard... shall specify a level of air quality the attainment and maintenance 

of which in the judgment of the Administrator, based on such criteria, is requisite to protect the 

public welfare from any known or anticipated adverse effects associated with the presence of 

such air pollutant in the ambient air.” The “criteria” referred to in this text are defined earlier in 

CAA section 108 (a) (2) which states in part that “[a]ir quality criteria for an air pollutant shall 

accurately reflect the latest scientific knowledge useful in indicating the kind and extent of all 

identifiable effects on public health or welfare which may be expected from the presence of such 

pollutant in the ambient air, in varying quantities.” Thus, while the criteria include “all 

identifiable effects”, Congress directed the EPA to establish the secondary NAAQS based on the 

Administrator’s judgement of what is requisite to protect against “adverse effects” in the context 

of the public welfare. However, the CAA does not provide specific standards for determining 

what constitutes an effect that is adverse to the public welfare leaving these determinations 

instead to the “judgment of the Administrator". As stated above in section 1.1, the PA is intended 

to help “bridge the gap” between the Agency’s scientific assessments presented in the ISA and 

REAs, which constitute “the criteria” and the judgments required of the EPA Administrator 

regarding whether it is appropriate to retain or revise the NAAQS.5

 In the context of the 

secondary standard, the PA thus serves the function of translating the information assessed in 

both the ISA and WREA into the public welfare policy context. In order to do this, the PA 

applies a specific approach or paradigm (see also section 1.3.2. above), that guides staff’s 

consideration and interpretation of the available information and which then informs staff 

conclusions regarding policy options that are appropriate for the Administrator to consider. The 

following discussion describes the evolution of this paradigm throughout the last several reviews 

and into the current review. 

In the 1997 secondary O3 NAAQS review, a policy-relevant distinction was made 

between the terms “injury” and “damage”. Specifically, O3-induced “injury” to vegetation was 

defined as encompassing all plant reactions, including reversible changes or changes in plant 

metabolism (e.g., altered photosynthetic rate), altered plant quality or reduced growth. In 

contrast, “damage” was defined to include only those injury effects that reach sufficient 

magnitude as to also reduce or impair the intended use or value of the plant, thus potentially 

being adverse to the public welfare. In published scientific literature, on the other hand, the terms 

“adverse”, “injury” or “damage” continue to be used interchangeably. The early O3 NAAQS 

reviews focused primarily on O3-induced effects at the individual and species level. In such 

 

5

American Farm Bureau Federation v. EPA, 559 F. 3d 512, 521 (D.C. Cir. 2009); Natural Resources 

Defense Council v. EPA, 902 F. 2d 962, 967-68, 970 (D.C. Cir. 1990). 
5-9 

cases, examples of vegetation effects that were also classified as damage included reductions in 

aesthetic values (e.g., visible foliar injury in ornamental species or occurring in valued natural 

landscapes such as national parks) and tree growth/biomass and crop yield losses (i.e., in terms 

of weight, number, quality, appearance, or size of harvestable crop or timber species). In the 

context of evaluating effects on single plants or species grown in monocultures such as managed 

forests, this construct continues to remain useful (73 FR 16492/96). 

 In subsequent reviews, however, the scientific literature linking O3 effects on plants or 

species to effects at the community or ecosystem level continued to increase. As a result, more 

recent reviews have considered a more expansive construct or paradigm of what appropriately 

constitutes O3 “damage” to extend beyond that of the individual or species level. A number of 

these broader paradigms have been discussed in the literature (72 FR 37890; Hogsett et al., 1997; 

Young and Sanzone, 2002). Thus, in the 2008 review, the Administrator, while continuing to 

express support for relying on a definition of “adverse” discussed in section IV.A.3 of the 

proposal (72 FR 37889-37890) that embeds “the concept of ‘intended use’ of the ecological 

receptors and resources that are affected”, also supported applying “that concept beyond the 

species level to the ecosystem level” (73 FR 16496). In so doing, the Administrator took note of 

“a number of actions taken by Congress to establish public lands that are set aside for specific 

uses that are intended to provide benefits to the public welfare, including lands that are to be 

protected so as to conserve the scenic value and the natural vegetation and wildlife within such 

areas, and to leave them unimpaired for the enjoyment of future generations” (73 FR 16496). 

Thus, this paradigm recognized that the significance to the public welfare of O3-induced effects 

on sensitive vegetation growing within the U.S. can vary depending on the nature of the effect, 

the intended use of the sensitive plants or ecosystems, and the types of environments in which 

the sensitive vegetation and ecosystems are located. Accordingly, any given O3-related effect on 

vegetation and ecosystems (e.g., biomass loss, crop yield loss, visible foliar injury) may be 

judged to have a different degree of impact on the public welfare depending, for example, on 

whether that effect occurs in a Class I area, a city park, or commercial cropland. In the 2010 

proposed reconsideration, the Administrator proposed to place the highest priority and 

significance on vegetation and ecosystem effects to sensitive species that are known to or are 

likely to occur in federally protected areas such as national parks and other Class I areas, or on 

lands set aside by states, tribes and public interest groups to provide similar benefits to the public 

welfare (75 FR 3023/24). Effects occurring in such areas would likely have the highest potential 

for being classified as adverse to the public welfare, due to the expectation that these areas need 

to be maintained in a more pristine condition to ensure their intended use is met. In contrast, in 

that proposal, the Administrator considered it less clear the degree to which O3 vegetation 

impacts potentially predicted to occur in areas and on species that are already heavily managed 
5-10 

to obtain a particular output (such as commodity crops or commercial timber production), would 

impair the intended use at a level that would be judged adverse to the public welfare and also 

noted that these species would likely receive some protection for a standard set to provide 

protection in areas set aside to be maintained in a more pristine condition (75 FR 3024). 

In the current review, we revisited the appropriateness of using this paradigm and 

whether the available information supported any further evolution. In so doing, we noted the 

ISA text, which states that “[o]n a broader scale, ecosystem services may provide indicators for 

ecological impacts. Ecosystem services are the benefits that people obtain from ecosystems 

(UNEP, 2003)” (U.S. EPA, 2013, Preamble, p. 1xxii) and the ISA list of a number of ecosystem 

services that can be affected by O3-induced effects on plants and ecosystems, including 

decreased productivity, decreased carbon sequestration, altered water cycling, and altered 

community composition (U.S. EPA, 2013, Figure 2-2, pp. 2-36; Figure 9-1, p. 9-3). We further 

noted that other recent EPA documents have already incorporated this concept. For example, the 

recent review of the secondary NAAQS for oxides of nitrogen and sulfur recognized that 

changes in ecosystem services may be used to aid in characterizing a known or anticipated 

adverse effect to public welfare and that an evaluation of adversity to the public welfare might 

consider the likelihood, type, magnitude, and spatial scale of the effect, as well as the potential 

for recovery and any uncertainties relating to these conditions (77 FR 20232). Similarly, the 

EPA document, Ecological Benefits Assessment Strategic Plan, includes a definition of 

ecological goods and services used for the purposes of benefits assessment that EPA has relied 

upon in regulatory impact analyses for previous rulemakings. This definition states that 

ecological goods and services are the ‘‘outputs of ecological functions or processes that directly 

or indirectly contribute to social welfare or have the potential to do so in the future”...and that 

“[s]ome outputs may be bought and sold, but most are not marketed’’ (U.S. EPA, 2006b). 

After considering this information, and given the accepted use of these concepts and their 

clear applicability to the secondary NAAQS review, we concluded that while it is still 

appropriate to apply the paradigm used in the 2010 reconsideration that takes into account the 

variation in public welfare significance of O3-related vegetation effects when evaluating the 

potential adversity of the currently available evidence, there is also sufficient support for an 

expansion of this paradigm to explicitly include consideration of impacts to ecosystem goods and 

services. Doing so can help clarify the relationship between predicted O3-induced vegetation 

effects and anticipated impacts on public welfare benefits received from those impacted species 

or ecosystems, and, as was done in the WREA, clarify how those services might be expected to 

change under air quality scenarios representing the current and potential alternative secondary 

standards (U.S. EPA, 2014a, chapter 5). The expansion of this paradigm to include ecosystem 
5-11 

goods and services brings with it a number of additional considerations. Specifically, when 

considering the potential public welfare benefits from these goods and services, it is important to 

note that they can accrue across a range of dimensions, including spatial, temporal, and social, 

and these likely will vary depending on the type of effect being characterized. For example, 

ecosystems can cover a range of spatial scales, and the services they provide can accrue locally 

or be distributed more broadly such as when crops are sold and eaten locally and/or also sold in 

regional, national and world markets. Ecosystem services can likewise be realized over a range 

of temporal scales from immediate up to long term (e.g., the removal of air pollutants that have a 

short-term impact on human health but are also climate forcers with long atmospheric lifetimes, 

such that their removal may have immediate as well as long-term benefits). The size of the 

societal unit receiving benefits from ecosystem services can also vary dramatically. For 

example, a national park can provide direct recreational services to the thousands of visitors that 

come each year, but also provide an indirect value to the millions who may not visit but receive 

satisfaction from knowing it exists and is preserved for the future (U.S. EPA, 2014a, chapter 5, 

section 5.5.1). 

We thus conclude that it is appropriate for the Administrator, in specifying what “level of 

air quality” for a pollutant “is requisite to protect the public welfare from any known or 

anticipated adverse effects associated with the presence of such air pollutant in the ambient air,” 

to evaluate the scientific evidence regarding these effects in the context of the most recent 

paradigm discussed above. This paradigm integrates the concepts of: 1) variability in public 

welfare significance given intended use and value of the affected entity such as an individual 

species; 2) relevance of associated ecosystem services to public welfare; 3) variability in spatial, 

temporal, and social distribution of ecosystem services associated with known and anticipated 

welfare effects. In so doing, we recognize that there is no bright-line rule delineating the set of 

conditions or scales at which known or anticipated effects become adverse to public welfare. 

Thus, the evidence and exposure/risk information discussed in this chapter will be further 

evaluated in Chapter 6 using the concepts incorporated in this paradigm to help inform the 

Administrator’s judgments with respect to the adversity of the effects to the public welfare and 

what is considered requisite protection. 

5.2 FOREST TREE GROWTH, PRODUCTIVITY AND CARBON STORAGE 

Trees merit consideration from a public welfare perspective because they provide many 

services that people value, including aesthetic value (also discussed in section 5.4 below), food, 

fiber, timber, other forest products, habitat, recreational opportunities, climate regulation, erosion 

control, air pollution removal, hydrologic and fire regime stabilization (U.S. EPA, 2014a, section 
5-12 

6.1, Figure 6-1, section 6.4, Table 6-13). One source identifies as many as 1,497 native tree 

species growing in the lower 48 of the U.S.6 Ozone has been shown to be phytotoxic to a number 

of important U.S. tree species with respect to growth, productivity, and carbon storage, including 

for cumulative exposures that have occurred under recent U.S. air quality. This section includes 

a discussion of the policy-relevant evidence and weight-of-evidence conclusions discussed in the 

ISA (section 5.2.1) and the exposure/risk results, including both quantitative and qualitative 

results for these effects, as well as associated ecosystem services (section 5.2.2) as described in 

the final WREA (U.S. EPA, 2014a). Important uncertainties and limitations in the available 

information are also discussed in each section. These discussions highlight the information we 

consider relevant to answering the overarching question and associated policy-relevant questions 

included in this section. 

5.2.1 Evidence-based Considerations 

 To what extent has scientific information become available that alters or 

substantiates our prior conclusions of O3-related effects on forest tree growth, 

productivity and carbon storage and of factors that influence associations between 

O3 concentrations and these effects? 

Research published since the 2006 CD substantiates prior conclusions regarding O3-

related effects on forest tree growth, productivity and carbon storage. The ISA states that 

“previous O3 AQCDs concluded that there is strong evidence that exposures to O3 decreases 

photosynthesis and growth in numerous plant species” and that “[s]tudies published since the 

2008 review support those conclusions” (U.S. EPA, 2013, p. 9-42). The recent studies that 

support the previous conclusions come from a variety of different study types that cover an array 

of different species, effects endpoints, levels of biological organization and exposure methods 

and durations. As stated in Chapter 1, and above, the documentation of O3-induced speciesspecific responses across multiple lines of evidence, and over the full range of levels of 

biological organization highlights and strengthens the consistency and coherence of the evidence 

available in this review. 

The previously available strong evidence for trees includes robust exposure-response (ER) functions for seedling biomass loss in 11 species developed under the National Health and 

Environmental Effects Research Laboratory-Western Ecology Division (NHEERL-WED) 

program. This series of experiments used open-top-chambers (OTC) to study seedling growth 

response for a single growing season under a variety of O3 exposures (ranging from near 

 

6

 USDA, NRCS. 2014. The PLANTS Database (http://plants.usda.gov, 7 July 2014). National Plant Data 

Team, Greensboro, NC 27401-4901 USA. 
5-13 

background to well above current ambient concentrations) and growing conditions (U.S. EPA 

2013, section 9.6.2, Lee and Hogsett, 1996). The evidence from these studies shows that there is 

a wide range in sensitivity across the studied species in the seedling growth stage over the course 

of a single growing season, with some species being extremely sensitive and others being very 

insensitive, or alternatively quite tolerant, over the range of cumulative O3 exposures studied 

(See Figure 5-1, below). 

In addition, field-based studies of species growing in natural stands have compared 

observed plant response across a number of different sites and/or years when exposed to varying 

ambient O3 exposure conditions only. For example, a study conducted in forest stands in the 

southern Appalachian Mountains found that the cumulative effects of ambient levels of O3

decreased seasonal stem growth (measured as a change in circumference) by 30-50% for most of 

the examined tree species (i.e., tulip poplar, black cherry, red maple, sugar maple) in a high O3

year in comparison to a low O3 year (McLaughlin et al., 2007a). The authors also reported that 

high ambient O3 concentrations can increase whole-tree water use and in turn reduce late-season 

streamflow (McLaughlin et al., 2007b) (U.S. EPA, 2013, p. 9-43). This study used ambient O3

conditions found at several different sites to create the variation in O3 exposures. 

Because trees and other perennials are long lived, it is important to consider the potential 

for impacts beyond a single year. Limited evidence in previous reviews reported that vegetation 

effects from a single year of exposure to elevated O3 could be observed in the following year. 

For example, growth affected by a reduction in carbohydrate storage in one year may result in 

the limitation of growth in the following year. Such “carry-over” effects have been documented 

in the growth of some tree seedlings and in roots (U.S. EPA, 2013, section 9.4.8; Andersen, et 

al., 1997). In the current review, additional field-based evidence expands our understanding of 

the consequences of single and multi-year O3 exposures in subsequent years. A number of 

studies were conducted at a planted forest at the Aspen Free-Air Carbon Dioxide Enrichment 

(FACE) site in Wisconsin. These studies, which occurred in a field setting more similar to 

natural forest stands than OTC studies, observed tree growth responses when grown in single or 

two species stands within 30-m diameter rings and exposed to ambient and above ambient 

conditions over a period of ten years. Some researchers similarly recognized the potential for 

carry-over effects when they observed that the effects of O3 on birch seeds (reduced weight, 

germination, and starch levels) could lead to a negative impact on species regeneration in 

subsequent years, and that the effect of reduced aspen bud size might have been related to the 

observed delay in spring leaf development. These effects suggest that elevated O3 exposures 

have the potential to alter carbon metabolism of overwintering buds, which may have subsequent 

effects in the following year (Darbah, et al., 2008, 2007; Riikonen et al., 2008; U.S. EPA, 2013, 

section 9.4.3). Other studies found that, in addition to affecting tree heights, diameters, and main 
5-14 

stem volumes in the aspen community, elevated O3 over a 7-year study period was reported to 

increase the rate of conversion from a mixed aspen-birch community to a community dominated 

by the more tolerant birch, leading the authors to conclude that elevated O3 may alter intra- and 

inter-species competition within a forest stand (Kubiske et al., 2006; Kubiske et al., 2007) (U.S. 

EPA, 2013, section 9.4.3). These studies confirm earlier FACE results showing large decreases 

in growth for aspen over a 6-7 year period when exposed to elevated O3 (King et al., 2005) and 

that yearly biomass loss cumulated over that timeframe. 

In addition to individual studies, recent meta-analyses have quantified the effect of O3 on 

trees across large numbers of studies. In particular, a recent meta-analysis (Wittig, et al., 2007) 

indicates a relationship between O3 concentrations in the northern hemisphere and stomatal 

conductance and photosynthesis, which decrease growth (U.S. EPA, 2013, section 9.4.3.1; 

Wittig et al., 2007). 7 This analysis reported that recent O3 concentrations in the northern 

hemisphere are decreasing stomatal conductance (13%) and photosynthesis (11%) across tree 

species. It also found that younger trees (<4 years) were affected less by O3 than older trees 

(Wittig, et al., 2007). A second meta-analysis, Wittig, et al. (2009), which quantitatively 

compiled peer-reviewed studies from the past 40 years, found that ambient O3 concentrations 

reported in those studies significantly decreased annual total biomass growth (7%) across the 263 

studies (U.S. EPA, 2013, section 9.4.3.1). The ISA states that this meta-analysis demonstrates 

the coherence of O3 effects across numerous studies and species that used a variety of 

experimental techniques, and these results support the conclusion of the previous CD that 

exposure to O3 decreases plant growth. Other meta-analyses have examined the effect of O3

exposure on root growth and generally found that O3 exposure reduced carbon allocated to roots. 

For example, Grantz et al. (2006) found that O3 exposure reduced the ratio between the relative 

growth rate of the root and shoot by 5.6% (U.S. EPA, 2013, pp. 9-45 to 9-46). 

In our consideration of the recent studies discussed above, in combination with the entire 

body of available evidence, we note that the recent scientific literature further strengthens and 

contributes to the consistency and coherence of the evidence base by substantiating and 

expanding prior conclusions regarding O3-related effects on tree growth, productivity and carbon 

storage, including mixed species forest stands and the ecosystems and services that derive from 

them, as discussed more fully below. We also note that the ISA concludes that the currently 

available evidence supports causal determinations regarding O3 effects on tree growth and 

productivity and the associated effects of altered carbon allocation to below ground tissues, rates 

of leaf and root production, turnover and decomposition that can alter below-ground 

 

7

 Meta-analysis allows for the objective development of a quantitative consensus of the effects of a 

treatment across a wide body of literature. 
5-15 

biogeochemical cycles, as well as the likely to be a causal relationship with reduced carbon 

sequestration and alteration of terrestrial community composition and water cycling (U.S. EPA, 

2013, Table 2-2; 9-19). Finally, we note that except for the recent limited information on 

cottonwood in the ISA (U.S. EPA, 2013, section 9.6.3.3), there has not been an expansion in the 

number of tree species for which we have E-R functions, so only 12 species have available E-R 

functions for use in quantitative exposure and risk analyses and for predicting tree seedling 

response under a range of O3 exposure conditions/scenarios. While these 12 species represent 

only a small fraction (0.8%) of the total number of native tree species in the contiguous U.S. 

(1,497), this small subset includes eastern and western species, deciduous and coniferous species, 

and species that grow in a variety of ecosystems and represent a range of tolerance to O3 (Figure 

5-1 below, U.S. EPA 2013, section 9.6.2; U.S. EPA, 2014a, section 6.2, Figure 6-2, Table 6-1). 

The CASAC states in their letter to the Administrator on the second draft PA, that while “[t]here 

is considerable uncertainty in extrapolating from the 12 forest tree species to all forest tree 

species in the U.S...[i]t is scientifically justifiable to extrapolate from the known E-R curves, 

assuming that they are representative of the un-sampled population” (Frey, 2014, p. 15). 

 As we further consider the results from the quantitative exposure and risk analyses, 

described below and in the WREA (U.S. EPA, 2014a), that in addition to the quantifiable portion 

of risks associated with the robust information on tree species, it is also reasonable to consider, 

based on the long-standing evidence and recent CASAC advice, the anticipated risks to other tree 

species that have not had their sensitivity to O3 studied in a robust quantifiable way but that 

potentially have O3 sensitivities that fall within the range for known species (see U.S. EPA, 

2007, Table 7J-1 in Appendix 7J and Table 7J-2). 

 To what extent have important uncertainties in the evidence identified in the last 

review been reduced and/or new uncertainties emerged? 

As stated above, the ISA concludes that the new evidence confirms, strengthens and 

expands our understanding of O3 effects on plants. Much of this new evidence is focused on the 

molecular and genetic level, providing important new mechanistic information that in some cases 

enhances our understanding of the complexity of the O3–plant response. This information has, in 

general, reduced overall uncertainties at the subcellular and cellular scales (U.S. EPA, 2013, 

section 9.3.6). 

Other recent information has also reduced some associated uncertainties regarding O3

impacts at the whole plant, species, and ecosystem scales. Importantly, one key uncertainty 

related to the potential broader applicability of OTC-generated tree seedling E-R functions to 

estimate biomass loss under different (i.e., field) O3 exposure conditions has been significantly 

reduced (U.S. EPA, 2013, section 9.6). Using recent field-based information available in the 
5-16 

current review, we conducted an analysis comparing OTC data with FACE data for one crop and 

one tree species (U.S. EPA, 2013, section 9.6.3.2). One comparison was done using soybean 

OTC data from the National Crop Loss Assessment Network (NCLAN)8 and more recent fieldbased data from the SoyFACE experiment, as discussed in section 5.3 below. The second was 

done using aspen seedling OTC data from the NHEERL-WED studies and more recent fieldbased data from the Aspen FACE study site. The result of the aspen analysis showed very close 

agreement between the biomass loss predictions based on OTC data and Aspen FACE 

observations, even when comparing the results of experiments that used different exposure 

methodologies, different genotypes, locations, and durations. The soybean analysis showed 

similar agreement between the OTC data and the SoyFACE experiment. Based on this analysis, 

the ISA concluded that “[o]verall, the studies at the Aspen FACE experiment were consistent 

with many of the open-top chamber (OTC) studies that were the foundation of previous O3

NAAQS reviews” and that “[t]hese {recent} results strengthen the understanding of O3 effects on 

forests and demonstrate the relevance of the knowledge gained from trees grown in OTC 

studies” (U.S. EPA 2013, p. 2-38, Section 9.6.3). The ISA additionally notes that with respect to 

aspen, “the function based on one year of growth was applicable to subsequent years” (of the sixyear dataset) (U.S. EPA, 2013, section 9.6.3.2). This result is significant in that it shows that at 

least for this species, the seedling E-R function was able to predict responses beyond the seedling 

growth stage. While recognizing that some uncertainties remain for E-R functions for some 

individual species for which the database is relatively less robust, taken together, this 

information substantially reduces uncertainties associated with use of the tree seedling OTCderived E-R functions to predict the response of tree seedlings in field settings and in some cases 

beyond the seedling growth stage. This information in combination with results from recent 

meta-analyses, as discussed above, reduces the uncertainties associated with potential impacts of 

other experimental factors on the O3-plant response. Thus, in the current review, we have greater 

confidence than in the last review in using these E-R functions to estimate tree growth response 

outside the chamber setting (U.S. EPA, 2013, section 9.6.2; U.S. EPA, 2014a, section 6.2). 

Several uncertainties are specific to studying or modeling O3 impacts on trees, and derive 

from the long lifespan of trees, which can range from decades to centuries. Because most studies 

are designed to take place within an annual or 2-3 year timeframe, typically information is 

available for only a small fraction of the lifetime of a tree. Given this reality, one uncertainty 

 

8

 The NCLAN program was conducted from 1980 to 1987 at five different locations across the US. At 

each site, open top chambers were used to expose plants to O3 treatments that represented the range of 

concentrations that occur in different areas of the world. The NCLAN focused on the most important U.S 

agricultural crops (Heagle et al., 1989; http://www.ars.usda.gov/Main/docs.htm?docid=12462). 
5-17 

that remains is the degree to which exposures in a single year or over multiple years affect trees 

over the longer term. However, as discussed above, recent studies from the Aspen FACE site 

have reduced this uncertainty by providing additional evidence that demonstrates that exposures 

in one year have the potential to cause effects in a subsequent year (carry-over effects) and that 

the annual effects from exposures over multiple years have the potential to compound (U.S. 

EPA, 2013, 9.4.3, pp. 9-42 to 9-47). Such effects, when they cumulate from one or more years 

of elevated O3 exposures, can lead to more serious longer-term impacts on growth, reproduction, 

recruitment, and competitive interactions within forest stands, and at larger spatial scales (U.S. 

EPA, 2013, p. 1-8), which would also have ramifications for any associated ecosystem services. 

In recognition of this recent evidence, the current CASAC Panel advised that “[a] 2% biomass 

loss is an appropriate scientifically based value to consider as a benchmark of adverse impact for 

long-lived perennial species such as trees, because effects are cumulative over multiple years” 

and stated that in its “scientific judgment, it is appropriate to identify a range of levels of 

alternative W126-based standards that includes levels that aim for not greater than 2% RBL for 

the median tree species” (Frey, 2014, p. 14). The CASAC further states that it “considers it 

significant that a similar value of 1% - 2% for tree seedling biomass loss was recommended 

previously by a consensus meeting of experts on ecological effects of ozone (Heck and Cowling, 

1997)” (Frey, 2014, p. 14). 

 A related uncertainty comes from the limited evidence showing that sensitivity to O3 can 

vary over the lifespan of trees and that this variation in growth-stage sensitivity is speciesspecific. For example, some species have been shown to be more sensitive during younger 

growth stages (i.e., seedling/sapling) while other species may be more sensitive as adults. 

Though a few studies have examined tree growth beyond the seedling stage (e.g., aspen) and in 

some species has been measured for both seedling and mature trees within a species (e.g., red 

oak), for most studied tree species it remains uncertain to what degree effects observed during 

one growth stage can be extrapolated to other growth stages. An analysis in the WREA 

comparing seedling to adult tree biomass loss, discussed in 5.2.2 below, informs our 

consideration of this remaining uncertainty (U.S. EPA, 2014a, section 6.2.1.1). 

These uncertainties are taken into account when we consider how much weight to put on 

predictions of risks for known effects and how precautionary it is appropriate to be in light of the 

potential for cumulative effects from multiple year exposures that could reasonably be 

anticipated to occur, based on the evidence above. 
5-18 

 To what extent does currently available evidence suggest locations where the 

vulnerability of sensitive species, ecosystems and/or their associated services to 

O3-related effects on tree growth, productivity and carbon storage would have 

special significance to the public welfare? 

A number of different types of locations provide services of special significance to the 

public welfare. These services can flow in part or entirely from the vegetation that grows there 

(see also discussion under section 5.1 above). With respect to forested lands, the WREA notes 

that there are approximately 751 million acres of forest lands in the U.S., one third of which (250 

million acres) is federally owned (U.S. EPA, 2014a, p. 5-15). In order to identify what types of 

forest locations have special significance from a public welfare perspective, it is first useful to 

consider the types of services that can flow from forested areas, and more specifically, from 

forested areas with trees that are sensitive to O3. Some sensitive tree species provide public 

welfare benefits based on their cultural significance, and some lands are important to the public 

welfare for their cultural value. For example, tribal lands, federally designated Class I areas, 

non-Class I national parks and wilderness areas, and other areas set aside to provide similar 

public welfare benefits, are valued for their cultural services such as outdoor recreation and 

aesthetics. Appendix 5A includes a table listing known O3-sensitive species, including some 

trees that have been identified as having cultural importance to some tribes (U.S. EPA, 2014a, 

section 6.4.2). Locations where these species are growing and are used by tribes to support 

cultural practices would thus be potentially vulnerable to impacts from elevated cumulative O3 

exposures, which could result in the loss of those associated cultural services, including those 

associated with sensitive tree species. Class I areas and other parks have also been afforded 

special federal protection to preserve services such as a healthy natural environment that 

provides for the enjoyment of these resources unimpaired for current and future generations, 

sustainable native plant and wildlife populations, and unique recreational opportunities. As 

mentioned above, 66 O3-sensitive species have been identified on NPS and U.S. Fish and 

Wildlife Service lands).9

 Other forested lands, both public and private, where trees are grown for 

timber production could also be at risk, especially in a single timber species stand that is 

sensitive to O3 (i.e., Ponderosa pine) (see WREA section 6.3 and section 5.2.2 below). Urban 

forests provide a number of important services to the public, such as air pollution removal, 

cooling of the heat island effect, and beautification (U.S. EPA, 2014a, section 6.6.2). These 

 

9

 See http://www2.nature.nps.gov/air/Pubs/pdf/flag/NPSozonesensppFLAG06.pdf 
5-19 

urban forests have also been recognized as important to environmental justice communities.10 

Because urban forests can include O3-sensitive trees (e.g., black cherry), O3 exposures have the 

potential to reduce the services they provide. The WREA analysis of five urban case study areas, 

discussed below, quantified the O3 impacts on air pollution removal and carbon sequestration in 

those urban areas (WREA, sections 6.6.2 and 6.7; section 5.2.2 below). Black cherry, for 

example, was one of the top ten occurring species in four of the five case study areas. 

The above types of forested lands have clearly designated purposes or intended uses that 

help define the types of services that might be recognized as important from a public welfare 

perspective. In addition, other services provided by trees are potentially extremely valuable, but 

limited information is available to quantitatively value the extent of these services. Perhaps one 

of the most significant of these ecosystem services is climate regulation, which provides 

widespread and long-lasting public welfare benefits that the ISA determined is likely being 

compromised by the phytotoxic effects of O3 on tree growth, productivity and carbon storage. 

By reducing the amount of carbon taken up by plants, more CO2 is allowed to remain in the 

atmosphere where it potentially exacerbates the effects of climate change. In contrast to the 

location-specific discussion of services above, this service is potentially important to the public 

welfare no matter in what location the sensitive trees are growing, or what their intended current 

or future use. In other words, the benefit exists as long as the tree is growing, regardless of what 

additional functions and services it provides. 

In addition to identifying forested locations that provide ecosystem services that are 

important to the public welfare, we must also consider to what extent there is the potential for O3

to affect sensitive tree species growing on those lands to a degree sufficient to affect the public 

welfare. In so doing we first note that not all tree species are equally sensitive to O3 and thus not 

equally vulnerable to current ambient O3 exposures or those anticipated under various air quality 

scenarios. In further considering the degree to which O3-induced impacts to ecosystem services 

associated with such trees might be expected to occur, we first focused on the 12 species of trees 

for which we have E-R functions. While all of these species provide goods and services that are 

important to the public welfare, not all species are equally sensitive to O3 under recent ambient 

exposure conditions or conditions projected for adjusted air quality. Table 5-1 below (modified 

from WREA Table 6-13), provides a more detailed description of the ecosystem services 

provided by each of these species that benefit the public welfare. For the purposes of this 

 

10 See http://www.fs.fed.us/research/urban/environmental-justice.php and Federal Interagency Working 

Group on Environmental Justice. (2011). Community-Based Federal Environmental Justice Resource Guide. 

August. Available at http://www.epa.gov/environmentaljustice/resources/publications/interagency/resourceguide.pdf
5-20 

discussion we have ordered the species in the table to go in descending order from most to least 

sensitive (based on their predicted relative biomass loss (RBL) at a W126 of 15 ppm-hrs). 

 

Table 5-1. O3-Sensitive Trees, Their Uses and Relative Sensitivity 

Tree Species O3 Effect Role in Ecosystems and Public Welfare Uses 

Eastern 

Cottonwood11

Populus 

deltoides

Biomass loss Containers, pulp, and plywood 

Erosion control and windbreaks 

Quick shade for recreation areas 

Beaver dams and food

Black Cherry 

Prunus serotina 

Biomass loss, 

Visible foliar 

injury 

Cabinets, furniture, paneling, veneers, crafts, toys 

Cough remedy, tonic, sedative 

Flavor for rum and brandy 

Wine making and jellies 

Food for song birds, game birds, and mammals

Eastern White 

Pine 

Pinus strobus 

Biomass loss Commercial timber, furniture, woodworking, and Christmas trees 

Medicinal uses as expectorant and antiseptic 

Food for song birds and mammals 

Used to stabilize strip mine soils

Quaking Aspen 

Populus 

tremuloides

Biomass loss, 

Visible foliar 

injury 

Commercial logging for pulp, flake-board, pallets, boxes, and plywood 

Products including matchsticks, tongue depressors, and ice cream sticks 

Valued for its white bark and brilliant fall color 

Important as a fire break 

Habitat for variety of wildlife 

Traditional native American use as a food source 

Yellow (Tulip) 

Poplar 

Liriodendron 

tulipifera

Biomass loss, 

Visible foliar 

injury 

Furniture stock, veneer, and pulpwood 

Street, shade, or ornamental tree – unusual flowers 

Food for wildlife 

Rapid growth for reforestation projects

Ponderosa Pine 

Pinus 

ponderosa 

Biomass loss, 

Visible foliar 

injury 

Lumber for cabinets and construction 

Ornamental and erosion control use 

Recreation areas 

Food for many bird species, including the red-winged blackbird, chickadee, 

finches, and nuthatches

 

11 The E-R function for cottonwood is considered less robust because it is based on the results of a single 

gradient study (Gregg et al., 2003).
5-21 

Tree Species O3 Effect Role in Ecosystems and Public Welfare Uses 

Red Alder 

Alnus rubra 

Biomass loss, 

Visible foliar 

injury 

Commercial use in products such as furniture, cabinets, and millwork 

Preferred for smoked salmon 

Dyes for baskets, hides, moccasins 

Medicinal use for rheumatic pain, diarrhea, stomach cramps – the bark 

contains salicin, a chemical similar to aspirin 

Roots used for baskets 

Food for mammals and birds – dam and lodge construction for beavers 

Conservation and erosion control

Red Maple^ 

Acer rubrum 

Biomass loss One of the most abundant and widespread in eastern U.S. Used for 

revegetation, especially in riparian buffers and landscaping, where it is 

valued for its brilliant fall foliage, some lumber and syrup production. 

Important wildlife browse food, especially for elk and white-tailed deer in 

winter, also leaves important food source for some species of butterflies and 

moths.

Virginia Pine 

Pinus 

virginiana

Biomass loss, 

Visible foliar 

injury 

Pulpwood, strip mine spoil banks and severely eroded soils 

Nesting for woodpeckers, food for songbirds and small mammals

Sugar Maple 

Acer saccharum 

Biomass loss Commercial syrup production 

Native Americans used sap as a candy, beverage – fresh or fermented into 

beer, soured into vinegar and used to cook meat 

Valued for its fall foliage and as an ornamental 

Commercial logging for furniture, flooring, paneling, and veneer 

Woodenware, musical instruments 

Food and habitat for many birds and mammals

Loblolly Pine* Biomass loss, 

visible foliar 

injury 

Most important and widely cultivated timber species in the southern U.S. 

Furniture, pulpwood, plywood, composite boards, posts, poles, pilings, 

crates, boxes, pallets. Also planted to stabilize eroded or damaged soils. It 

can be used for shade or ornamental trees, as well as bark mulch. 

Provides habitat, food and cover for white-tailed deer, gray squirrel, fox 

squirrel, bobwhite quail and wild turkey, red-cockaded woodpeckers, and a 

variety of other birds and small mammals. Standing dead trees are 

frequently used for cavity nests by woodpeckers. 

Douglas Fir 

Pseudotsuga 

menziesii

Biomass loss Commercial timber 

Medicinal uses, spiritual and cultural uses for several Native American 

tribes 

Spotted owl habitat 

Food for mammals including antelope and mountain sheep 

*Sensitivity categories added by EPA staff but not based on official designations.

Sources: USDA-NRCS, 2013; Burns, 1990; Hall and Braham, 1998. ^Red maple information from 

http://www.na.fs.fed.us/pubs/silvics_manual/volume_2/acer/rubrum.htm. *Loblolly pine use information from 

http://www.ncsu.edu./project/dendrology/index/plantae/vascular/seedplants/gymnosperms/conifers/pine/pinus/austra

les/loblollypine/html. 
5-22 

While we recognized that there are important ecosystem services provided by those 

species that are less sensitive to O3, those species would likely receive less benefit from 

additional protection below the current standard. In contrast, the other species would likely see 

improvements in their associated ecosystem services, some significantly, from an improvement 

in air quality. However, at the highest end of the known sensitivity spectrum, there are different 

issues that must be considered when evaluating the usefulness of this information in answering 

the above questions. The E-R function that is available for cottonwood is based on the results of 

a single gradient study (Gregg et al., 2003) and is considered less robust than the other E-R 

functions developed in OTCs. That combined with its apparent extreme response to O3

prompted CASAC to advise the Administrator to not place too much emphasis on cottonwood in 

the review of the secondary standard (Frey, 2014, p. 10). As a result, we have decided it would 

not be appropriate to use the cottonwood biomass loss estimates when considering what levels of 

W126 should be considered protective of median species biomass loss (see Table 5C-3). 

However, in this discussion of ecosystem services, we believe it is important to include 

cottonwood, given the many ecosystem services cottonwood provides (see Table 5-1 above), and 

several unique features that potentially make it and its associated ecosystem goods and services 

particularly vulnerable to impacts from O3. Specifically we note that cottonwood: 1) is often 

found growing along streams in riparian zones under well watered conditions that make it more 

susceptible to injury than species growing in areas that experience drier conditions in 

conjunction with higher O3 exposures; 2) can be the only tree species growing in certain types of 

ecosystems, thus providing important habitat for some organisms; 3) is fast growing and used 

commercially for pulpwood, manufacturing furniture and as a possible source for energy 

biomass (Burns and Hankola, 1990); 4) has provided limited, though still uncorroborated, 

evidence of the potential for the existence of extremely sensitive plant species which can 

reasonably be anticipated to exist and that could be impacted at similar cumulative exposures. 

With regard to the latter, we observe that CASAC also expressed the view that it “should be 

anticipated that there are species of vegetation that are highly sensitive to ozone that do not have 

E-R functions, and others that are insensitive. It is scientifically justifiable to extrapolate from 

the known E-R curves, assuming that they are representative of the un-sampled population” 

(Frey, 2014, p. 16). We also note that upon revisiting the available literature in the ISA 

following CASAC’s review of the second draft WREA and PA, we found two studies on a 

related European species (Populus nigra) that showed that this species had an O3 sensitivity that 

appears similar in magnitude to the U.S. cottonwood (Populus deltoides) based on its response 

for other growth endpoints as compared with the response of the other study species (Bortier, et 

al., 2000; Novak, et al., 2007) (U.S. EPA, 2006, AX9, pp. 91, 240; U.S. EPA, 2014a, Table 6-5). 
5-23 

This additional limited evidence of Populus seedling/sapling growth response, though not 

directly comparable to the U.S. study (Gregg et al., 2003) with respect to species, exposure 

methods, measurement endpoints and exposure values, does, in our judgment, lend some support 

to the observed magnitude of the reported U.S. cottonwood response. 

In addition to the information provided here on these 12 species, we note that there are 

many other species of trees with known or suspected O3-sensitive vegetation, such as those 

included in the 66 species identified on NPS and US Fish and Wildlife Service lands),12 species 

used in the USFS biomonitoring network, and various ornamental and agricultural species (i.e., 

Christmas trees, fruit and nut trees) that currently provide ecosystem services important to the 

public welfare, but whose vulnerability to impacts from O3 on tree growth, productivity and 

carbon storage has not been sufficiently characterized to allow it to directly inform our 

quantitative assessments (U.S. EPA, 2014a, Chapter 6; Abt Associates, 1995). However, as 

noted by CASAC, the anticipated impacts on these and other unstudied species should not be 

ignored or assumed insignificant. It is more likely that the range of O3 sensitivities found in the 

studies tree species likely reflects the range of O3 sensitivities in all tree species. 

Other factors that should be taken into account when considering the potential degree to 

which O3 might affect the ecosystem service flows from forested ecosystems are 1) the type of 

stand or community in which the sensitive species is found (i.e., single species versus mixed 

canopy); 2) the role or position the species has in the stand (i.e., dominant, sub-dominant, 

canopy, understory); 3) the O3 sensitivity of the other co-occurring species (O3 sensitive or 

tolerant); 4) environmental factors (drought or well watered conditions, other stressors). 

In light of the above discussion, it is clear that there are numerous locations where the 

vulnerability of O3-sensitive tree species to impacts from O3 on tree growth, productivity and 

carbon storage and their associated ecosystems and services could have special significance to 

the public welfare. Confirmation that the American public values healthy forests is provided in 

the WREA, which shows that Americans are willing to pay to protect forests from the damaging 

effects of air pollutants (U.S. EPA, 2014a, Chapter 5, pp. 5-16). Data provided by the National 

Survey on Recreation and the Environment (NSRE) indicates that Americans have very strong 

preferences for the non-use values of existence, bequest, and option services related to forests. 

Studies (Haefele et al., 1991, Holmes and Kramer, 1995) assess willingness-to-pay (WTP) for 

spruce-fir forest protection in the southeastern U.S. from air pollution and insect damage and 

confirm that the non-use values held by the survey respondents were in fact greater than the use 

or recreation values. The results of this survey showed that median household WTP was 

 

12 See http://www2.nature.nps.gov/air/Pubs/pdf/flag/NPSozonesensppFLAG06.pdf 
5-24 

estimated to be roughly $29 (in 2007 dollars) for the minimal protection program and $44 for the 

more extensive program. After decomposing their value for the extensive program into use, 

bequest, and existence values, the results were 13 percent for use value, 30 percent for bequest, 

and 57 percent for existence value (See U.S. EPA, 2014a, Table 5-6). These services may be at 

risk in areas where O3-sensitive trees are found. 

 To what extent does the available evidence indicate the occurrence of O3-related 

effects on forest growth, productivity and carbon storage attributable to 

cumulative exposures lower than previously established or that might be expected 

to occur under the current standard? 

The evidence base available in this review, as in the previous review, indicates that O3-

induced effects on tree growth, productivity and carbon storage can occur across a range of 

cumulative exposures, including those lower than previously established and that would be 

expected to occur under the current standard. In reaching this determination, we first consider 

the 11 tree seedling species for which robust E-R functions have been developed from the 

extensive evidence base of O3–induced growth effects that was also available and relied upon in 

the previous review. Each of these species were studied in OTCs, with most species studied 

multiple times under a wide range of exposure and/or growing conditions, with separate E-R 

functions developed for each species/exposure condition/growing condition scenario 

combination or case. Using all the information available from these multiple study cases (52 

cases in all), a robust composite E-R function was developed for all species combined and 

separate individual composite functions were derived for each species using cases that were 

available on individual species. These species-specific composite E-R functions have been 

successfully used to predict tree seedling species biomass loss response over a range of 

cumulative exposure conditions. Figure 5-1A, B below, which includes the 11 robust composite 

E-R functions available in the last review and the E-R for cottonwood (also described in U.S. 

EPA 2013, section 9.6.2 and U.S. EPA, 2014a, section 6.2, Table 6-1 and Figure 6-2), illustrates 

the appreciable variability in sensitivity that exists across the 12 studied species, and shows that 

biomass loss can occur over a wide range of cumulative exposures, including those previously 

established. This figure further shows that for some species biomass loss would be predicted to 

occur at very low cumulative exposures that can occur under air quality conditions that meet or 

are below the current standard (see Table 5-2 below). While we put less emphasis on 

cottonwood (as explained above), we do note that in answering the question above, it does 

provide limited recent evidence of the potential for effects of a greater magnitude and at lower 

cumulative exposures to occur than those considered in the last review and at exposures that 

would be allowed by the current standard. To the extent that such effects could be anticipated, 
5-25 

the cumulative exposures that could be allowed by the current secondary standard would not be 

protective. 

Figure 5-1. A) Relative biomass loss in seedlings for 12 studied species using composite 

functions in response to seasonal O3 concentrations in terms of seasonal W126 index 

values, Y-axis scale for RBL values represents 0% up to 100% (U.S. EPA 2014a, Figure 

6-2). 

0 10 20 30 40 50

0.0 0.2 0.4 0.6 0.8 1.0

W126

RBL

Red Maple

Sugar Maple

Red Alder

Tulip Poplar

Ponderosa Pine

White Pine

Loblolly Pine

Virginia Pine

Cottonwood

Aspen

Black Cherry

Douglas Fir
5-26 

 B) Expanded view of relative biomass loss in seedlings for 12 studied species using 

composite functions in response to seasonal O3 concentrations in terms of lower range of 

seasonal W126 index values, Y-axis scale for RBL values represents 0% up to 10% (U.S. 

EPA 2014a, Figure 6-2). 

In further answering the question above, we note CASAC’s advice that a 6% median 

RBL is unacceptably high, and that the 2% median RBL is an important benchmark to consider. 

Based on the information above, the median RBL is at or below 2% at the lowest W126 level 

assessed, 7 ppm-hrs. As the W126 level is incrementally increased, median RBL also increases 

incrementally, so that at W126 index values of 9, 11, 13, 15, 17, 19 and 21, the median RBL 

increases to 2.4%, 3.1%, 3.8%, 4.5%, 5.3%, 6.0% and 6.8%, respectively. Based on air quality 

analyses of 2009-2011 (Appendix 2B), there are approximately 342, 199, 92, 43, 24, 9, 3 and 0 

monitors with 3-year average W126 index values above 7, 9, 11, 13, 15, 17, 19 and 21 ppm-hrs 

when meeting the current standard. We note that these counts of monitors are based on those 

meeting the current standard and that there are many monitors for the 2009-2011 period that do 

not meet the current standard and also are above the W126 values of 7-21 ppm-hrs. 

 We also consider it informative to examine the individual species responses and RBL 

over the same W126 range. We first note, based on Figure 5-1(B) above that over the range of 7 

to 17 ppm-hrs, 5 species maintain RBLs of less than 2%. These more tolerant species include 
5-27 

Douglas fir, loblolly pine, Virginia pine, sugar maple and red maple. Two of these species (red 

maple and sugar maple) are estimated to have RBL levels above 2% at a W126 of 21. Black 

cherry, the most sensitive of the remaining six species, has RBL ranging from 35.57% at W126 

of 17 down to 16.67% at the W126 index value of 7 ppm-hrs. 

Additional evidence of the potential for O3-induced effects on tree seedling growth, 

productivity and carbon storage occurring under air quality scenarios allowed by the current 

standard is shown in Table 5-2 below. Specifically, all monitor sites in Table 5-2 have 3-year 8-

hour average values that meet the current standard, ranging from 67 to 75 ppb, have 3-year 

average W126 index values that are above 15 ppm-hrs, and are located in Class I areas. Across 

these 22 Class I areas, the highest single-year W126 index values for these three-year periods 

ranged from 17.4 to 29.0 ppm-hrs. In 20 of the areas, distributed across eight states (AZ, CA, 

CO, KY, NM, SD, UT, WY) and four regions (west, southwest, west/north central and central), 

this range was 19.1 to 29.0 ppm-hrs, exposure values for which the corresponding median 

species RBL estimates equal or exceed 6%, which CASAC termed “unacceptably high”. In 

addition, given that other environmental factors can influence the extent to which O3 may have 

the impact predicted by the E-R functions in any given year, we also note that the highest three 

year periods, that include these highest annual values for the 21 areas, are at or above 19 ppmhrs, ranging up to 22.5 ppm-hrs (which the median species RBL estimate is above 7%). 

Additionally, the highest three-year average W126 index value for each of the 22 areas (during 

periods meeting the current standard) was at or above 19 (ranging up to 22.5 ppm-hrs) in 11 

areas, distributed among five states in the west and southwest regions (U.S. EPA, 2014c, Table 

5-2, Appendix 5B). 

In addition, as data permit, Table 5-2 shows the studied tree species that are found in each 

of these Class I areas. Quaking aspen and ponderosa pine are two tree species that are found in 

many of these 22 parks and have a sensitivity to O3 exposure that places them near the middle of 

the group for which E-R functions have been established. In the areas where ponderosa pine is 

present, the highest single year W127 index values ranged from 18.7 to 29.0 and the highest 3-

year average W126 values in which these single year values are represented ranged from 15 to 

22.5, with these three-year values above 19 ppm-hrs in eight areas across five states. The 

ponderosa pine RBL estimates for 29 and 22.5 ppm-hrs are approximately 12% and 9%, 

respectively. In areas where quaking aspen is present, the highest single year W127 index values 

ranged from 19.2 to 26.7 ppm-hrs and the highest 3-year average W126 values in which these 

single year values are represented ranged from 15.0 to 22.2, with values above 19 ppm-hrs in 

eight areas across five states. The quaking aspen RBL estimates for 26.7 and 22.2 ppm-hrs are 

approximately 16% and 13%, respectively. Based on this, we note growth effects associated with 
5-28 

exposure concentrations occurring during periods where the current standard is met in many of 

these Class I areas. On the basis of such information, Table 5-2 provides evidence of the 

potential for significant growth loss in locations where ambient conditions meet the current 

standard. 

Table 5-2. O3 concentrations in Class I areas during period from 1998 to 2012 that met 

the current standard and where three-year average W126 index value was at or above 

15 ppm-hrs.* 

Class I Area State / 

County 

Design 

Value 

(ppb)* 

3-year Average 

W126 

(ppm-hrs)* 

(# ≥ 19 ppm-hrs, 

range)

Annual W126 

(ppm-hrs)* 

(# ≥ 19 ppm-hrs, range)

Number of 

3-year 

Periods 

Bandelier Wilderness Area 

QA, DF, PP

NM / 

Sandoval 70-74 15.8-20.8 (2, 20.0-20.8) 12.1-25.3 (4, 19.2-25.3) 8 

Bridger Wilderness Area 

QA, DF WY / Sublette 69-72 15.1-17.4 9.9-19.2 (1, 19.2) 5 

Canyonlands National Park 

QA, DF, PP, UT / San Juan 69-73 15.0-20.5 (2, 19.8-

20.5) 9.9-24.8 (5, 19.3-24.8) 9 

Carlsbad Caverns National 

Park PP NM / Eddy 69 15.0-15.3 8.6-26.7 (1, 26.7) 3 

Chiricahua National 

Monument DF, PP AZ / Cochise 69-73 15.7-18.0 13.2-21.6 (2, 19.3-21.6) 7 

Grand Canyon National 

Park QA, DF, PP

AZ / 

Coconino 68-74 15.6-22.2 (7, 19.2-22.3) 11.3-26.7 (7, 19.8-26.7) 12 

John Muir Wilderness Area 

QA, DF, PP CA / Inyo 71-72 16.5-18.6 10.1-25.8 (2, 23.9-25.8) 3 

Lassen Volcanic National 

Park DF, PP CA / Shasta 75 15.3 13.6-18.7 (1, 18.7) 1 

Mammoth Cave National 

Park BC, C, LP, RM, SM, VP, YP 

KY / 

Edmonson 74 15.9 12.5-22.5 (1, 22.5) 1 

Mesa Verde National Park

DF

CO / 

Montezuma 67-73 15.5-21.0 (2, 19.0-21.0) 10.7-23.6 (4, 19.7-23.6) 10 

Mokelumne Wilderness 

Area DF, PP CA / Amador 74 17.6 14.8-22.6 (1, 22.6) 1 

Petrified Forest National 

Park AZ / Navajo 70 15.7 12.9-19.2 (1, 19.2) 1 

Pinnacles National 

Monument

CA / San 

Benito 74 15.1 13.1-17.4 1 

Rocky Mountain National 

Park QA, DF, PP

CO / Boulder 73-75 15.1-19.3 (1, 19.3) 9.5-25.1 (5, 20.7-25.1) 6 

 CO / 

Larimer 74 15.0-18.3 11.1-25.8 (3, 19.1-25.8) 3 

Saguaro National Park DF, PP AZ / Pima 69-74 15.4-18.9 11.0-23.1 (3, 20.0-23.1) 6 

Sierra Ancha Wilderness

Area DF, PP AZ / Gila 72-75 17.9-22.4 (3, 20.2-22.4) 14.8-27.5 (4, 20.3-27.5) 4 

Superstition Wilderness 

Area PP

AZ / 

Maricopa 75 22.4 (1, 22.4) 14.5-28.6 (2, 27.4-28.6) 1 
5-29 

AZ / Pinal 73-75 18.7-22.5 (2, 20.9-

22.5) 14.8-29.0 (3, 22.6-29.0) 3 

Weminuche Wilderness 

Area QA, DF, PP CO / La Plata 70-74 15.0-19.1 (1, 19.1) 10.9-21.0 (2, 20.8-21.0) 5 

West Elk Wilderness Area 

QA, DF

CO / 

Gunnison 68-73 15.6-20.1 (1, 20.1) 12.9-23.9 (3, 21.1-23.9) 8 

Wind Cave National Park 

QA, PP SD / Custer 70 15.4 12.2-20.6 (1, 20.6) 1 

Yosemite National Park QA, 

DF, PP

CA / 

Tuolumne 73-74 20.7-20.8 (2, 20.7-

20.8) 19.7-22.1 (4, 19.7-22.1) 2 

Zion National Park QA, DF, PP UT / 

Washington 70-73 17.8-21.1 (2, 20.3-21.1) 14.9-24.2 (5, 19.3-24.2) 4 

*Based on data from http://www.epa.gov/ttn/airs/airsaqs/detaildata/downloadaqsdata.htm (US EPA, 2014c). W126 

values are truncated after first decimal place. 

Superscript letters refer to species present for which E-R functions have been developed. 

QA=Quaking Aspen, BC=Black Cherry, C=Cottonwood, DF=Douglas Fir, LP=Loblolly Pine, PP=Ponderosa Pine, 

RM=Red Maple, SM=Sugar Maple, VP=Virginia Pine, YP=Yellow (Tulip) Poplar. 

Sources for presence of species include U.S. Department of Agriculture databases in 2014 

http://www.fs.fed.us/foresthealth/technology/nidrm2012.shtml, http://plants.usda.gov, 

http://www.wilderness.net/printFactSheet.cfm?WID=583 

 In answering the above question, we note that less information is available from fieldbased studies (e.g., FACE, gradient) due to the absence of robust E-R functions, the limited 

range of exposure scenarios evaluated, and unavailability of study exposures in terms of daily 8-

hour averages. 

Taken together, the information described above provides consistent and coherent 

evidence that O3-induced impacts on tree seedling growth, productivity and carbon storage are 

occurring at cumulative exposures allowed by the current standard. In particular, this information 

provides clear evidence of the potential for significant growth loss in Class I locations where 

ambient conditions meet the current standard. 

5.2.2 Exposure/Risk-based Considerations 

The WREA presents a number of quantitative analyses of exposure and risk related to 

tree growth, productivity and carbon storage intended to inform our consideration of exposure 

and risk associated with the current and potential alternative standards (Table 5-3 below; U.S. 

EPA, 2014a, Chapter 6). 
5-30 

Table 5-3. Exposure, risk and ecosystem services analyses related to tree growth, 

productivity and carbon storage. 

Species Level Effects Ecosystem Level Effects Ecosystem Services 

WREA 

estimatesA

Derivation of median biomass 

loss values from individual 

species E-R functions 

Comparison of tree seeding 

growth to that of mature trees 

Percent of total geographic 

areaB with annual relative 

biomass loss above 2% 

Number of assessed Class I 

areas with annual relative 

biomass loss above 2% 

 Economic surplus to timber 

producers and consumers 

(WREA, Table 6-12)

 Carbon storage, nationally 

(WREA, Table 6-19) 

 Carbon storage, in 5 urban 

areas (WREA, Table 6-21) 

 Air pollutant removal in 5 

urban areas (WREA, Table 

6-22) 

A See WREA Chapter 6 (U.S. EPA, 2014a). 

B The total geographic area includes only the contiguous U.S. 

The relevant quantitative exposure and risk analyses for tree biomass loss, productivity 

and carbon storage include: 

1) Species-specific and median biomass loss estimates from composite functions. 

2) National-scale assessments for: a) basal area weighted relative biomass loss for tree 

seedlings; b) timber production; c) carbon sequestration. 

3) Case study-scale assessments for: a) carbon sequestration; b) air pollution removal. 

 For what air quality scenarios were exposures and risks estimated? What 

approaches were used to estimate W126 exposures for those conditions? What 

are associated limitations and uncertainties? 

Quantitative exposure and risk analyses were conducted to evaluate the effects on tree 

growth, productivity and carbon storage, and associated ecosystem services, that would be 

predicted under five air quality scenarios (recent ambient, just meeting the current standard, and 

W126 potential alternative standards of 15, 11, and 7 ppm-hrs). Table 5-5 summarizes the 

methodology used to develop the quantitative estimates for each of the five air quality scenarios. 

In general, this methodology involved two steps. The first is derivation of the average W126 

index value (across the three years) at each monitor location. This value is based on unadjusted 

data for recent conditions and adjusted concentrations for the four other scenarios. The 

development of adjusted concentrations was done for each of 9 regions independently (see U.S. 

EPA, 2014a, section 4.3.4.1). In the second step, national-scale spatial surfaces (W126 index 

values for each 12 x 12 km2

 grid cell from the Community Multi-scale Air Quality (CMAQ) 

model) were created using the monitor-location values and the Voronoi Neighbor Averaging 

(VNA) spatial interpolation technique (details on the VNA technique are presented in U.S. EPA, 

2014a, Appendix 4A). 
5-31 

Table 5-4. Summary of methodology by which national surface of 3-year average W126 

index values was derived for each air quality scenario. 

Scenario

 Development of W126 index values for Each Air Quality Scenario 

Monitor-location-specific calculations 

and any model-based adjustment 

Derivation of national surface of 

average W126 index values

Recent 

Conditions 

(2006-2008) 

An annual W126 index value is calculated for each year at 

each monitor location, using the highest 3-month period. A 

location-specific 3-year W126 was calculated by averaging 

annual W126 index values from 3 consecutive years which 

may have used different 3-month periods. 

The VNA method is applied to the 

monitor-location average W126 

index values to create a national 

distribution of W126 index values 

within model grid-cells for each 

scenario. 

Current 

Standard 

2006-2008 hourly O3 concentrations at each monitor location 

are adjustedA to create a three year record of O3

concentrations that just meets the current standard (see 

WREA, section 4.3.4). This results in air quality at other 

monitors well below the level of the controlling monitor. 

A seasonal W126 index value is calculated for each year at 

each monitor location using the same 3-month period for 

each year (which is the highest as a 3-yr average and is 

highest in at least one of the years). A location-specific 

average is derived from these three index values. 

Average 

W126 Index 

of 15 ppm-hrs 

First, hourly O3 concentrations were adjusted to just meet the 

current standard. Second, hourly O3 concentrations at each 

monitor location, within each modeling region, are adjusted 

to create a record for which the highest location-specific 

average index value in the region (the controlling location) 

just meets the scenario target index value. 

A seasonal W126 index value is calculated for each year (of 

2006-2008 period) at each monitor location, using the same 

3-month period for each year (which is the highest in at least 

one of the years). A location-specific average is derived 

from these three index values. 

Average 

W126 Index 

of 11 ppm-hrs 

Average 

W126 Index 

of 7 ppm-hrs 

A The model-based adjustment approach is based on regional emission reduction scenarios at monitor sites 

followed by spatial interpolation for broader spatial coverage. See WREA, chapters 3 and 4, and Appendix 4A. 

During the recent conditions period (2006 through 2008), the average W126 index values 

(across the three-year recent conditions period) at the monitor locations ranged from below 5 

ppm-hrs to 48.6 ppm-hrs (U.S. EPA 2014a, Figure 4-4 and Table 4-3). Across the nine modeling 

regions, the maximum average W126 index values ranged from 48.6 ppm-hrs in the west region 

down to 6.6 ppm-hrs in the northwest region. After adjusting the 2006-2008 data to just meet the 

current standard in each region, the region-specific maximum values range from 18.9 ppm-hrs in 

the west region to 2.6 ppm-hrs in the northeast region (U.S. EPA, 2014a, Table 4-3). After 

application of the VNA technique to the current standard scenario monitor location values, the 

average W126 index values were below 15 ppm-hrs across the national surface with the 
5-32 

exception of a very small area of the southwest region (near Phoenix) where the average W126 

index values was near or just above 15 ppm-hrs. Thus, it can be seen that application of the 

interpolation method to estimate W126 index values at the centroid of every 12 x 12 km2 grid 

cell rather than only at each monitor location results in a lowering of the highest values. 

 What are the nature and magnitude of exposure- and risk-related estimates for 

tree growth, productivity, and carbon storage under recent conditions or 

conditions remaining upon meeting the current standard? To what extent are 

these exposures and risks important from a public welfare perspective? 

In answering the above question, the WREA performed a number of different 

assessments to estimate the exposures and risks predicted under the five air quality scenarios 

across a range of spatial scales. These assessments include those for individual species response 

as well as the median species response for studied species ranging from the county scale up to 

estimations of exposures and risks to ecosystem services associated with forests at the urban, 

park and national scales. 

Before conducting the exposure and risk assessments, the WREA examined three 

approaches for characterizing the median response, as shown in Figure 5-2 below (U.S. EPA, 

2014a, section 6.2.1.2 and Figure 6-5). These approaches use the 11 robust E-R functions for 

tree seedlings from the OTC research and the cottonwood E-R function. For some species, only 

one study was available (e.g., red maple), and for other species there were as many as 11 studies 

available (e.g., ponderosa pine). The first approach plotted the median (red line) of all 52 tree 

seedling studies available (across the 12 species). In this first approach, species with multiple 

studies would be represented more than once in the median. The second approach characterized 

the median (green line) by combining the composite E-R functions, when available for species 

with multiple studies, with the E-R functions for species with a single study available13 for each 

of the 12 tree species. In this second approach, each species is represented only once in the 

median. The third approach used a stochastic sampling method to randomly select a single E-R 

function from the studies available for each of the 12 species. The process was repeated 1,000 

times (grey lines), and the median value was plotted for biomass loss values of 1% to 7%, and 

10% (red dots; the bar associated with each median point denotes the 25th and 75th percentile 

values). This third approach illustrates the effect of within-species variability on estimates of the 

median response. The median W126 index values are similar when using the first two 

approaches; however, the median value is higher when within-species variability is included 

(U.S. EPA, 2014a, section 6.2.1.2). Across these three approaches, the median seasonal W126 

index value for which a two percent biomass loss is estimated in seedlings for the studied species 

 

13 For some species, only one study was conducted so that E-R function was used. 
5-33 

ranges between approximately 7 and 14 ppm-hrs. Using the green line, the seasonal W126 index 

value for which a two percent biomass loss is estimated in seedlings for the median of the 

composite functions for the 12 studied species is approximately 7 ppm-hrs. After reviewing these 

three approaches, the CASAC stated “[t]he Monte Carlo analysis (red dots, Figure 5-2) should 

not be used in evaluating the effect of ozone on RBL of tree seedlings. This analysis 

overemphasizes the species for which relatively few E-R functions are available, is biased 

toward the few less sensitive response functions available for some individual species, makes 

unsupported assumptions regarding the representativeness of available response functions, and 

confounds intra- and inter-species variability in unquantifiable ways. We favor using a measure 

of central tendency of the data, specifically the median across species (the green line in Figure 5-

2). This analysis provides the median of best available estimates within each species, and the 

median across species with all species treated equally” (Frey, 2014, p. 14). Given this advice, in 

selecting an approach for use in later analyses, we have chosen to use the green line because the 

approach that generated it incorporates all the information in a way that gives equal weight to 

each studied species without losing any of the available data. 

Figure 5-2. Relationship of tree seedling percent biomass loss with seasonal W126 index. 

(From U.S. EPA 2014a, Figure 6-5) 

The WREA used the E-R functions for 12 species described above with information on 

the distribution of those species across the U.S., and average W126 exposure estimates to 
5-34 

estimate relative biomass loss for each of the studied species for each national air quality 

scenario (U.S. EPA, 2014a, section 6.2.1.3 and Appendix 6A). For example, the estimates of 

relative biomass loss of ponderosa pine for air quality adjusted to just meet the current standard 

are illustrated in Figure 5-3 below. While relative biomass loss below 2% is estimated for most 

areas where this species is found, estimates in some areas of the southwest fall above 2% 

biomass loss (U.S. EPA 2014a, Figure 6-8 and Appendix 6A). 

Figure 5-3. Relative biomass loss of Ponderosa Pine for air quality adjusted to just meet 

the current standard (U.S. EPA 2014a, Figure 6-8). 

The WREA also developed national-scale estimates of O3 biomass ecosystem-level 

impacts considering the 12 studied species together (U.S. EPA 2014a, section 6.8, Table 6-25). 

This was done using the species-specific biomass loss E-R functions, information on prevalence 

of the studied species across the U.S., and a weighting approach based on proportion of the basal 

area within each grid cell that each species contributes. The RBL values for multiple tree species 

were weighted by their basal area and combined into a weighted RBL value (wRBL). The 

wRBL is intended to inform our understanding of the potential magnitude of the ecological effect 

that could occur in some ecosystems. Specifically, the more basal area that is affected in a given 
5-35 

ecosystem, the larger the potential ecological effect. A wRBL value for each grid cell is 

generated by weighting the RBL value for each studied tree species found within that grid cell by 

the proportion of basal area it contributes to the total basal area of all tree species within the grid 

cell, and then summing those individual wRBLs. The percent of total basal area that exceeds a 

2% weighted relative biomass loss in the recent conditions scenario is 10.1% (U.S. EPA 2014a, 

Table 6-25). Based on the average W126 index values estimated for the air quality scenario just 

meeting the current standard across the contiguous U.S., the WREA estimates 0.2% of the total 

geographic area to have a wRBL above 2% based on the E-R functions for the 11 tree species 

and 0.8% based on 12 tree species (U.S. EPA 2014a, Table 6-25). We recognize that these 

estimates are likely biased low as there may be other unstudied O3-sensitive tree species in some 

areas that are also being impacted at those levels. Further, this analysis does not take into 

account the effects of competition, which could further increase biomass loss in O3-sensitive 

species. 

In addition, the WREA characterized the number of counties where there would be one or 

more studied tree species showed a 2% biomass loss (U.S. EPA, 2014a, Table 6-7), which is 

shown in Table 5-5 below. This is consistent with CASAC advice that “rather than focusing 

solely on the median relative biomass loss (RBL), the number of counties containing sensitive 

tree species that are expected to have growth loss of greater than 2% should be quantified.” 

(Frey, 2014, p. 11). These data are presented as the number of U.S. counties in which any of the 

12 studied tree species exceeds 2% RBL, further categorized by the number of studied species 

that exceed that benchmark for each of the five air quality scenarios using 3-year average W126 

index values. In addition, this table provides the total number of counties (out of 3,109 total 

counties) for each exposure scenario with at least one species exceeding 2% RBL and the 

number of counties where the median of the composite functions for each species exceeds 2% 

RBL. The maximum number of species that exceed 2% RBL in any one county is five species, 

which only occurs under recent O3 conditions. After meeting the current standard, the maximum 

number of species in any one county is four. Because cottonwood and black cherry are highly 

sensitive species and to provide a reference for the effect of these species, the data are also 

presented excluding cottonwood and excluding cottonwood and black cherry. 

This information shows that a number of counties have more than one O3-sensitive 

species growing in it, potentially together in the same forest stands, whose RBLs are above 2%. 

Under recent conditions, the proportion of total counties of 3,109, with 1 or more species with an 

RBL greater than 2% is 89% (2,761 counties) for the scenario inclusive of cottonwood and black 

cherry. When air quality is adjusted to just meet the current standard, that proportion dropped to 

74% (2,313 counties). When air quality is adjusted to just meet a 3-year average W126 index 
5-36 

value of 15, 11 and 7 ppm-hrs, the proportion is 73%, 72% and 71%, respectively. For median 

RBL values, under recent conditions, 72% of the counties have median RBLs above 2%. When 

air quality is adjusted to the current standard, that proportion drops to 22% and further decreases 

to 20% for air quality adjusted to just meet a 3-year average W126 level of 7 ppm-hrs. 

Given CASAC’s advice to put less emphasis on cottonwood, we focus on the rows of this 

table that excluded cottonwood. Under recent air quality conditions, the proportion of counties 

with 1 or more species with an RBL greater than 2% is 78% (2,418 counties). As air quality is 

adjusted to just meet the current standard and the alternative W126 index value of 7 ppm-hrs, 

this number drops to 62% and 58%, respectively. In addition, under recent conditions, 52% of 

the counties have median RBLs above 2%. When air quality is adjusted to the current standard, 

that proportion drops to 8% and further decreases to 6% for air quality adjusted to just meet a 3-

year average W126 level of 7 ppm-hrs. 

 Table 5-5 also provides information on the influence of black cherry on the estimates 

and shows that black cherry is a very sensitive species that is widespread in the Eastern U.S. We 

note that of the 1,929 counties estimated to have 1 or more species with an RBL greater than 2% 

when meeting the current standard, 1,805 of those counties are estimated to have black cherry as 

the only specie estimated to experience this level of biomass loss. With respect to median RBL 

values, of the 239 counties estimated to have a median RBL above 2% when meeting the current 

standard, 203 of those counties have a RBL above 2% because of the presence of black cherry. 
5-37 

Table 5-5. Number of Counties with Tree Species Exceeding 2% Relative Biomass Loss. 

Number of species 

exceeding 2% RBL 

Number of Counties (3,109 Total) 

Recent 

Conditions 75 ppb 15 ppm-hrs 11 ppm-hrs 7 ppm-hrs 

All 12 tree species with E-R functions 

5 134 - - - - 

4 387 3 3 - - 

3 765 24 22 14 5 

2 882 994 981 972 924 

1 593 1,292 1,273 1,238 1,277 

0 348 796 830 885 903 

Total counties exceeding 2,761 2,313 2,279 2,224 2,206 

Counties exceeding for 

the median species 2,237 685 670 651 627 

11 tree species with E-R functions excluding cottonwood 

5 15 - - - - 

4 180 - - - - 

3 680 3 3 - - 

2 933 46 32 14 5 

1 610 1,880 1,857 1,818 1,812 

0 691 1,180 1,217 1,277 1,292 

Total counties exceeding 2,418 1,929 1,892 1,832 1,817 

Counties exceeding for 

the median species 1,604 239 221 204 172 

10 tree species with E-R functions excluding cottonwood and black cherry 

5 - - - - - 

4 15 - - - - 

3 187 - - - - 

2 856 29 15 2 1 

1 920 95 72 19 8 

0 1,131 2,985 3,022 3,088 3,100 

Total counties exceeding 1,978 124 87 21 9 

Counties exceeding for 

the median species 666 36 18 6 2 

We also consider WREA estimates (quantitative and qualitative) of effects on several 

ecosystem services. First, impacts on growth related to cumulative O3 exposure values in 

federally designated Class I areas were derived from an average wRBL value (discussed above) 

for 145 of the 156 Class I areas (U.S. EPA 2014a, section 6.8.1). Given established objectives 

for Class I areas (e.g., to maintain in perpetuity), effects in Class I areas may be considered to 
5-38 

have the potential to adversely affect the intended use of the ecosystem, e.g., to leave them 

unimpaired and preserve them for the enjoyment of future generations. Under recent conditions, 

this analysis estimates that 13 Class I areas have wRBL values above 2%. Further, this analysis 

estimates that based on average W126 index values estimated for the air quality scenario just 

meeting the current standard, 2 of the 145 Class I areas assessed would be expected to have 

multiple-species, wRBL values above 2% (U.S. EPA 2014a, Table 6-26). However, we 

recognize that this analysis is limited to the 12 studied tree species, and therefore could 

underestimate other O3-sensitive species without E-R functions. 

The WREA also presents national-scale estimates of the effects of biomass loss on timber 

production and agricultural harvesting, as well as on carbon sequestration. The WREA used the 

O3 E-R functions for tree seedlings to calculate relative yield loss (equivalent to biomass loss) 

across the trees’ entire life spans. Because the forestry and agriculture sectors are related and 

trade-offs occur between the sectors, the WREA also calculated the resulting market-based 

welfare effects of O3 exposure in the forestry and agriculture sectors.14 In the analyses for 

commercial timber production, based on the 3-year average W126 index values estimated for the 

air quality scenario just meeting the existing standard, RYL estimates for timber were below one 

percent with the exception of the Southwest, Southeast, Central, and South regions (U.S. EPA, 

2014a, section 6.3, Table 6-9) (see U.S. EPA, 2014a, Table 6-8 for clarification on region 

names). At the current standard the highest yield loss occurs in upland hardwood forests in the 

South Central and Southeast regions at over 3% per year and in Corn Belt hardwoods at just over 

2% loss per year. Relative yield losses for timber remain above one percent for the 3-year 

average W126 scenarios for 15 and 11 ppm-hrs in parts of the Southeast, Central, and South 

regions, and for the 7 ppm-hrs scenario in the Southeast and South regions (U.S. EPA, 2014a, 

section 6.3, Table 6-9). In addition, relative yield losses for timber were above two percent in 

parts of the Southeast and Central U.S. after just meeting the existing standard as well as in the 

15 ppm-hrs and 11 ppm-hrs scenario. 

In addition to estimating changes in forestry and agricultural yields, the WREA presents 

estimated changes in consumer and producer/farmer surplus associated with the change in yields. 

Changes in biomass affect individual tree species, but the overall effect on forest ecosystem 

productivity depends on the composition of forest stands and the relative sensitivity of trees 

 

14 The WREA used the Forest and Agricultural Sector Optimization Model with Greenhouse Gases 

(FASOMGHG). FASOMGHG is a national-scale model that provides a complete representation of the U.S. forest 

and agricultural sectors’ impacts of meeting alternative standards. FASOMGHG simulates the allocation of land 

over time to competing activities in both the forest and agricultural sectors. FASOMGHG results include multiperiod, multi-commodity results over 60 to 100 years in 5-year time intervals when running the combined forestagriculture version of the model. 
5-39 

within those stands. Economic welfare impacts resulting from just meeting the existing and 

alternative standards were largely similar between the forestry and agricultural sectors -- 

consumer surplus, or consumer gains, generally increased in both sectors because higher 

productivity under lower O3 concentrations increased total yields and reduced market prices. 

Comparisons are not straightforward to interpret due to market dynamics. For example, because 

demand for most forestry and agricultural commodities is not highly responsive to changes in 

price, there were more examples for which producer surplus (i.e., producer gains) declines. 15 In 

some cases, lower prices reduce producer gains more than can be offset by higher yields. The 

increase in consumer welfare is much larger than the loss of producer welfare resulting in net 

welfare gains in the forestry sector nationally. The national-scale analysis of carbon dioxide 

(CO2) sequestration estimates more storage under the current standard compared to recent 

conditions (U.S. EPA 2014a, Appendix 6B, Table B-10). In considering the significance of the 

potential climate and ecosystem service impact, we also note the large uncertainties associated 

with this analysis (see U.S. EPA 2014a, Table 6-27). 

We additionally consider the WREA estimates of tree growth and ecosystem services 

provided by urban trees over a 25-year period for five urban areas based on case-study scale 

analyses that quantified the effects of biomass loss on carbon sequestration and pollution 

removal (U.S. EPA 2014a, sections 6.6.2 and 6.7).16 The urban areas included in this analysis 

represent diverse geography in the Northeast, Southeast, and Central regions, although they do 

not include an urban area in the western U.S. Estimates of the effects of O3-related biomass loss 

on carbon sequestration indicate the potential for an increase of somewhat more than a million 

metric tons of CO2 equivalents for average W126 index values associated with meeting for the 

current standard scenario as compared to recent conditions. Somewhat smaller additional 

increases are estimated for the three W126 scenarios in comparison to the current standard 

scenario (U.S. EPA 2014a, section 6.6.2 and Appendix 6D). 

In addition to the quantitative assessments discussed above, qualitative assessments for 

some ecosystem services, were also conducted, such as commercial non-timber forest products 

and recreation (U.S. EPA, 2014a, section 6.4), aesthetic and non-use values (U.S. EPA, 2014a, 

section 6.4), increased susceptibility to insect attack and fire damage (U.S. EPA, 2014a, sections 

5.3 and 5.4, respectively). Other ecological effects that are causally or likely causally associated 

with O3 exposure such as terrestrial productivity, water cycle, biogeochemical cycle, and 

 

15 See Chapter 6, Section 6.3 of the WREA for a discussion of economic welfare and consumer and 

producer surplus. 

16 The WREA used the i-Tree model for the urban case studies. i-Tree is a peer-reviewed suite of software 

tools provided by USFS. 
5-40 

community composition (U.S. EPA 2013, Table 9-19) were not directly addressed in the WREA 

due to a lack of sufficient quantitative information. 

There is substantial heterogeneity in plant responses to O3, both within species, between 

species, and across regions of the U.S. The O3-sensitive tree species are different in the eastern 

and western U.S. -- the eastern U.S. has far more species. Ozone exposure and risk is somewhat 

easier to assess in the eastern U.S. because of the availability of more data and the greater 

number of species to analyze. In addition, there are more O3 monitors in the eastern U.S. but 

fewer national parks (U.S. EPA, 2014a, chapter 8). 

 What are the uncertainties associated with both quantitative and qualitative 

information?

Several key limitations and uncertainties, which may have a large impact on both overall 

confidence and confidence in individual analyses, are discussed here. Despite these 

uncertainties, the overall body of scientific evidence underlying the ecological effects and 

associated ecosystem services evaluated in the WREA is strong, and the methods used to 

quantify associated risks are scientifically sound (Frey, 2014). Key uncertainties associated with 

the assessment of impacts on ecosystem services at the national and case-study scales, as well as 

across species, U.S. geographic regions and future years include those associated with the 

interpolated and adjusted O3 concentrations used to estimate W126 exposures in the WREA air 

quality scenarios and those associated with the available seedling E-R functions. 

The WREA identifies sources of uncertainty for the W126 estimates for each air quality 

scenario and qualitatively characterizes the magnitude of uncertainty and potential for directional 

bias (U.S. EPA, 2014a, Table 4-5). These sources of uncertainty are described in more detail in 

the WREA Chapter 4 and summarized below. 

An important large uncertainty in the analyses is the assumed response of the W126 

concentrations to emissions reductions needed to meet the existing standard (U.S. EPA, 2014a, 

section 8.5.1). We note that any approach to characterizing O3 over broad geographic areas based 

on concentrations at monitor locations will convey inherent uncertainty. The model-based 

adjustments, based on U.S.-wide emissions reductions in oxides of nitrogen (NOx), do not 

represent air quality distributions from an optimized control scenario that just meets the current 

standard (or target W126 index values for other scenarios), but rather characterize one potential 

distribution of air quality across a region when all monitor locations meet the standard (U.S. 

EPA 2014a, section 4.3.4.2).17 An additional uncertainty comes from the creation of a national 

 

17 Because our analyses used U.S.-wide NOx emissions reductions to simulate just meeting the existing 

standard independently in each region, there are broad regional reductions in O3 even in meeting standards in urban 

areas when targeting a few high-O3 urban monitors for reductions. However, the assumption of broad regional or 

national NOx reductions are not unreasonable given EPA regulations such as the NOx SIP Call program 
5-41 

W126 surface using the VNA technique to interpolate recent air quality measurements of O3. In 

general, spatial interpolation techniques perform better in areas where the O3 monitoring network 

is denser. Therefore, the W126 estimated in the rural areas in the West, Northwest, Southwest, 

and West North Central with few or no monitors (Figure 2-1) are more uncertain than those 

estimated for areas with denser monitoring. Additionally, the surface is created from the threeyear average at the monitor locations, rather than creating a surface for each year and then 

averaging across years at each grid cell; the potential impact of this on the resultant estimates is 

considered in the WREA (U.S. EPA, 2014a, Appendix 4A). 

 Because the W126 estimates generated in the air quality analyses are inputs to the 

vegetation risk analyses for biomass loss, any uncertainties in the air quality analyses are 

propagated into the those analyses (U.S. EPA 2014a, section 8.5). In its letter to the 

Administrator following its review of the second draft WREA CASAC notes that:

“The currently reported finding of only small differences in risk between just meeting the 

current standard and a W126-based level of 15 ppm-hrs must not be interpreted to mean 

that just meeting the current standard will be as protective as meeting a W126-based 

standard at 15 ppm-hrs. There are two key factors that must be considered when making 

this comparison. First, air quality was simulated in the Second Draft WREA based on the 

magnitude of across-the-board reductions in NOx emissions required to bring the highest 

monitor down to the target level. Meeting a target level at the highest monitor requires 

substantial reductions below the targeted level through the rest of the region. This 

artificial simulation does not represent an actual control strategy and may conflate 

differences in control strategies required to meet different standards and different targets. 

As a result, there may be a number of monitors that meet the current standard but would 

not meet an alternative W126 standard. Second, and equally important, the current form 

of the standard is much less biologically relevant for protecting vegetation than is a 

seasonal, peak weighted index such as the W126, which was designed to measure the 

cumulative effects of ozone exposure.” (Frey, 2014, pp. 11-12). 

With regard to the robust seedling E-R functions, the description of Figure 5-2 above 

provides some characterization of the variability of individual study results and the impact of 

that on estimates of W126 index values that might elicit different percentages of biomass loss in 

tree seedlings (U.S. EPA, 2014a, section 6.2.1.2). Even though the evidence shows that there 

are additional species adversely affected by O3-related biomass loss, the WREA only has E-R 

functions available to quantify this loss for 12 tree species. This absence of information only 

 implemented to help areas meet the 1997 O3 standard resulting in substantial reductions in power plant NOx 

emissions from states across the eastern U.S., and the multitude of onroad and offroad mobile source rules that will 

lead to reduction in NOx from these sources across the country in future years. 
5-42 

allows a partial characterization of the O3-related biomass loss impacts in trees associated with 

recent O3 index values and with just meeting the existing and potential alternative secondary 

standards. In addition, there are uncertainties inherent in these E-R functions, including the 

extrapolation of relative biomass loss rates from tree seedlings to adult trees and information 

regarding within-species variability. The overall confidence in the E-R function varies by 

species based on the number of studies available for that species. Some species have low 

within-species variability (e.g., many agricultural crops) and high seedling/adult comparability 

(e.g., aspen), while other species do not (e.g., black cherry). The uncertainties in the E-R 

functions for biomass loss and in the air quality analyses are propagated into the analysis of the 

impact of biomass loss on ecosystem services, including provisioning and regulating services 

(U.S. EPA, 2014a, Table 6-27). The WREA characterizes the direction of potential influence of 

E-R function uncertainty as unknown, yet its magnitude as high, concluding that further studies 

are needed to determine how accurately the assessed species reflect the larger suite of O3-

sensitive tree species in the U.S. (U.S. EPA, 2014a, Table 6-27). 

Another uncertainty associated with interpretation of the WREA biomass loss-related 

estimates concerns the potential for underestimation of compounding of growth effects across 

multiple years of varying concentrations. Though tree biomass loss impacts were estimated 

using air quality scenarios of 3-year average W126 index values, the WREA also conducted an 

analysis to compare the impact of using a variable compounding rate based on yearly variations 

in W126 exposures to that of using a W126 index value averaged across three years. The WREA 

compared the compounded values for each region, except for the South. In these examples, one 

species was chosen that occurred within that region. Air quality values associated with just 

meeting the existing standard of 75 ppb were used. Within each region the WREA analysis used 

both the W126 index value at each monitor in the region for each year and the three-year average 

W126 index value using the method described in Chapter 4. The results show that the use of the 

three-year average W126 index value may underestimate RBL values slightly. However, it 

should be noted that the approach does not account for moisture levels or other environmental 

factors that could affect biomass loss (U.S. EPA, 2014a, section 6.2.1.4 and Figure 6-14). In 

considering these results, we note that in these regions and in all three years, the three-year 

average W126 index value is sometimes above and sometimes below the individual year W126 

index value. 

In the national-scale analyses of timber production, agricultural harvesting, and carbon 

sequestration, the WREA used the FASOMGHG model, which includes functions for carbon 

sequestration, assumptions regarding proxy species, and non-W126 E-R functions for three 

crops. However, FASOMGHG does not include agriculture and forestry on public lands, 

changes in exports due to O3 into international trade projections, or forest adaptation. Despite 
5-43 

the inherent limitations and uncertainties, the WREA concludes that the FASOMGHG model 

reflects reasonable and appropriate assumptions for a national-scale assessment of changes in 

the agricultural and forestry sectors due to changes in vegetation biomass associated with O3

exposure (U.S. EPA, 2014a, sections 6.3, 6.5, 6.6, and 8.5.2). 

In the case study analyses of five urban areas, the WREA used the i-Tree model, which 

includes an urban tree inventory for each area and species-specific pollution removal and carbon 

sequestration functions. However, i-Tree does not account for the potential additional VOC 

emissions from tree growth, which could contribute to O3 formation. Despite the inherent 

limitations and uncertainties, the WREA concludes that the i-Tree model reflects reasonable and 

appropriate assumptions for a case study assessment of pollution removal and carbon 

sequestration for changes in biomass associated with O3 exposure (U.S. EPA, 2014a, sections 

6.6.2, 6.7, and 8.5.2). 

The overall effect of the combined set of uncertainties on confidence in the interpretation 

of the WREA results is difficult to quantify. Due to differences in available information, the 

degree to which each analysis was able to incorporate quantitative assessments of uncertainty 

differed. Despite these uncertainties, the overall body of scientific evidence underlying the 

ecological effects and associated ecosystem services evaluated in the WREA is strong, and the 

methods used to quantify associated risks are scientifically sound (Frey, 2014). 

5.3 CROP YIELD LOSS 

This section considers the current evidence and exposure/risk information to inform 

consideration of the adequacy of the protection provided by the current standard from known and 

anticipated adverse welfare effects of O3 related to crop yield and other associated effects. Crops 

warrant consideration from a public welfare perspective because they provide food and fiber 

services to humans. This section includes a discussion of the policy-relevant science and weightof-evidence conclusions discussed in the ISA (section 5.3.1) and the exposure/risk results 

(section 5.3.2) described in the final WREA. Important uncertainties and limitations in the 

available information are discussed under the related question below. These discussions 

highlight the information we consider relevant to answering the overarching question and 

associated policy-relevant questions included in this section. 

5.3.1 Evidence-based Considerations 

Ozone can interfere with carbon gain (photosynthesis) and allocation of carbon. As a 

result of decreased carbohydrate availability, fewer carbohydrates are available for plant growth, 

reproduction, and/or yield. For seed-bearing plants, these reproductive effects will culminate in 
5-44 

reduced seed production or yield. The detrimental effect of O3 on crop production has been 

recognized since the 1960s, and current O3 concentrations in many areas across the U.S. are high 

enough to cause yield loss in a variety of agricultural crops including, but not limited to, 

soybeans, wheat, cotton, potatoes, watermelons, beans, turnips, onions, lettuces, and tomatoes. 

Increases in O3 concentration may further decrease yield in these sensitive crops while also 

causing yield losses in less sensitive crops (U.S. EPA 2013, section 9.4.4). The ISA concluded 

that the evidence is sufficient to determine that there is a causal relationship between O3

exposure and reduced yield and quality of agricultural crops (U.S. EPA 2013, Table 2-2). 

 To what extent has scientific information become available that alters or 

substantiates our prior conclusions regarding O3-related crop yield loss and of 

factors that influence associations between O3 levels and crop yield loss? 

In general, the vast majority of the new scientific information has substantiated our prior 

conclusions regarding O3 crop yield loss. On the whole, this evidence supports previous 

conclusions that exposure to O3 decreases growth and yield of crops. The ISA describes average 

yield loss reported across a number of meta-analytic studies have been published recently for 

soybean wheat, rice, semi-natural vegetation, potato, bean and barley (U.S. EPA 2013, section 

9.4.4.1). Further, several new exposure studies continue to show decreasing yield and biomass in 

a variety of crops with increased O3 exposure (U.S. EPA 2013, section 9.4.4.1, Table 9-17). 

Research has linked increasing O3 concentration to decreased photosynthetic rates and 

accelerated aging (U.S. EPA 2013, section 9.4.4) in leaves, which are related to yield. Recent 

research has highlighted the effects of O3 on crop quality. Increasing O3 concentration can also 

decrease nutritive quality of grasses and macro- and micro-nutrient concentrations in fruits and 

vegetable crops (U.S. EPA 2013, section 9.4.4). The findings of these studies did not change our 

understanding of O3-related crop loss since the last review and little information has emerged on 

factors that influence associations between O3 levels and crop yield loss. 

 To what extent have important uncertainties identified in the last review been 

reduced and/or new uncertainties emerged? 

Important uncertainties have been reduced regarding crop E-R functions, especially for 

soybean. In general, the ISA reports consistent results across exposure estimation techniques 

and across crop varieties. Figure 5-4 below illustrates the composite E-R functions for the 10 

crop species assessed in the WREA (U.S. EPA, 2014a, Figure 6-3). 
5-45 

 

Figure 5-4. Relative yield loss in crops using the composite functions for 10 studied 

species in response to seasonal O3 concentrations in terms of seasonal W126 index 

values, Y-axis scale for RYL values represents 0% up to 100% (U.S. EPA 2014a, Figure 

6-3). 

Two important uncertainties have been reduced regarding the E-R functions for yield 

effects of O3 in crop species, especially for soybean. First, in the last several reviews, the extent 

to which E-R functions developed in OTC predicted plant responses in the field and under 

different exposure conditions was not clear. In this review, the ISA included an analysis 

comparing OTC data with field-based data for one crop and one tree species (U.S. EPA, 2013, 

section 9.6.3.2). The crop comparison was done using soybean OTC data from NCLAN and 

field-based data from SoyFACE. The NCLAN program, which was undertaken in the early to 

mid-1980s, assessed multiple U.S. crops, locations, and O3 exposure levels, using consistent 

methods, to provide the largest, most uniform database on the effects of O3 on agricultural crop 

yields (U.S. EPA 1996; U.S. EPA 2006; U.S. EPA 2013, sections 9.2, 9.4, and 9.6, Frey, 2014, 

0 10 20 30 40 50

0.0 0.2 0.4 0.6 0.8 1.0

W126 (ppm-hrs)

RYL

Barley

Field Corn

Cotton

Kidney Bean

Lettuce

Peanut

Potato

Grain Sorghum

Soybean

Winter Wheat
5-46 

p. 9).18 The SoyFACE experiment was a chamberless field-based exposure study in Illinois that 

was conducted from 2001 – 2009 (U.S. EPA 2013, section 9.2.4). Yield loss in soybean from O3

exposure at the SoyFACE field experiment was reliably predicted by soybean E-R functions 

developed in NCLAN (U.S. EPA, 2013, Section 9.6). This analysis supports the robustness and 

use of the E-R functions developed in NCLAN to predict relative yield loss to O3 exposure in a 

realistic agricultural setting. 

A second area of uncertainty that was reduced is that regarding the appropriateness of 

applying the NCLAN E-R functions to more recent cultivars that are currently being grown. 

Because recent studies continue to find yield loss levels in crop species studied previously under 

NCLAN that reflect the earlier findings, the ISA concluded that there has been little new 

evidence that crops are becoming more tolerant of O3 (U.S. EPA, 2006a; U.S. EPA 2013). This 

is especially evident in the research on soybean. In a meta-analysis of 53 studies, Morgan et al. 

(2003) found consistent deleterious effects of O3 exposures on soybean from studies published 

between 1973 and 2001. Further, Betzelberger et al. (2010) recently utilized the SoyFACE 

facility to compare the impact of elevated O3 concentrations across 10 soybean cultivars to 

investigate intraspecific variability of the O3 response. The E-R functions derived for these 10 

current cultivars were similar to the response functions derived from the NCLAN studies 

(Heagle, 1989), suggesting there has not been any selection for increased tolerance to O3 in more 

recent cultivars. The 2013 ISA reported comparisons between yield predictions based on data 

from cultivars used in NCLAN studies, and yield data for modern cultivars from SoyFACE (U.S. 

EPA, 2013, section 9.6.3). They confirm that the average response of soybean yield to O3

exposure has not changed in current cultivars. Thus, staff concludes that at least for soybean, 

uncertainties associated with use of the NCLAN generated E-R functions to estimate biomass 

loss in recent cultivars has been reduced. 

 To what extent does the available evidence indicate the occurrence of O3-related 

effects on crop yield loss attributable to cumulative exposures at lower ambient O3

concentrations than previously established or to exposures at or below the level of 

the current standard? 

Little scientific evidence has emerged to indicate a lower W126 index value for 

cumulative exposures that can affect crop yield than previously established. However, as 

 

18 The NCLAN protocol was designed to produce crop exposure-response data representative of the areas 

in the U.S. where the crops were typically grown. In total, 15 species (e.g., corn, soybean, winter wheat, tobacco, 

sorghum, cotton, barley, peanuts, dry beans, potato, lettuce, turnip, and hay [alfalfa, clover, and fescue]), accounting 

for greater than 85 percent of U.S. agricultural acreage planted at that time, were studied. Of these 15 species, 13 

species including 38 different cultivars were combined in 54 cases representing unique combinations of cultivars, 

sites, water regimes, and exposure conditions. Crops were grown under typical farm conditions and exposed in 

open-top chambers to ambient O3, sub-ambient O3, and above ambient O3. 
5-47 

discussed below, CASAC has provided a target benchmark protection for crop yield loss that can 

help better focus a discussion of the level of exposure that O3 related effects on crops can occur 

(levels of concern). Currently available evidence supports effects on crop yield at cumulative 

exposures at and below the level of the current standard. As described above, the new evidence 

has strengthened the basis for using the information from the E-R functions. 

Based on the 10 robust E-R functions (i.e., barley, lettuce, field corn, grain sorghum, 

peanut, winter wheat, cotton, soybean, potato and kidney bean) described in the ISA and 

additionally analyzed in the WREA (Figure 5-4), Table 5C-3 shows that for the CASAC 

recommended target benchmark protection level of 5% for median crop relative yield loss 

(RYL), W126 index values ranging from 7 to 17 ppm-hrs are protective. However, when 

individual species are considered over this same range, the proportion of crops protected varies 

from 5/10, 6/10, 7/10, 9/10, 10/10, and 10/10 at the W126 levels of 17, 15, 13, 11, 9, and 7 ppmhrs. To the extent a given species is judged as having particular importance to the public 

welfare, breaking the information down by species can be helpful. For example, less than 5% 

yield loss was estimated for soybeans at the W126 index value of 12 ppm-hrs (U.S. EPA 2014a, 

Figure 6-3). Four of the studied crop species (barley, lettuce, field corn, and grain sorghum) are 

more tolerant, with RYL under 1% over the W126 range from 7 to 17 ppm-hrs. Peanut also 

remained under 4% RYL over the same W126 range. Other species differed regarding the W126 

level at which RYL reached or fell below 5%. Specifically, for winter wheat, cotton, soybean, 

kidney bean and potato, the relevant W126 index values at which RYLs were below 5% are 15, 

13, 11, 11, and 9 ppm-hrs. 

Where the current evidence on crop yield loss is not in terms of parts per billion 

concentrations over a specific exposure period such as eight hours, assessing whether O3

concentrations associated with meeting the current standard would allow crop yield effects is 

more complex. In order to characterize the O3 exposures associated with crop yield loss in terms 

of seasonal W126 index and to consider the extent to which such index values might be expected 

to occur in agricultural locations that meet the current standard, we evaluated two agricultural 

counties in Kansas using O3 monitoring data from EPA’s Air Quality System (AQS) combined 

with the E-R function for soybeans. Sedgwick and Sumner counties both met the level of the 3-

year, 8-hour standard of 75 ppb in 2009-2011, but both still had a maximum annual W126 level 

of 19 ppm-hours in 2011. At that annual W126 index value, soybean yield loss would be 

predicted to be 9% in those counties in that year. 
5-48 

 To what extent does currently available evidence suggest locations where the 

vulnerability of sensitive species, ecosystems and/or their associated services to 

O3-related crop yield loss would have special significance to the public welfare? 

During the previous NAAQS reviews, there were very few studies that estimated O3

impacts on crop yields at large geographical scales (i.e., regional, national or global). Recent 

modeling studies of the historical impact of O3 concentrations found that increased O3 generally 

reduced crop yield, but the impacts varied across regions and crop species (U.S. EPA, 2013, 

Section 9.4.4.1). The largest O3-induced crop yield losses were estimated to occur in highproduction areas exposed to elevated O3 concentrations, such as the Midwest and the Mississippi 

Valley regions of the United States. Among crop species, the estimated yield loss for wheat and 

soybean were higher than rice and maize. Additionally, satellite and ground-based O3

measurements have been used to assess yield loss caused by O3 over the continuous tri-state area 

of Illinois, Iowa, and Wisconsin. The results indicate that O3 concentrations during the assessed 

period reduced soybean yield, which correlates well with the previous results from FACE- and 

OTC-type experiments (U.S. EPA 2013, section 9.4.4.1). 

Thus, the recent scientific literature in the ISA continues to support the conclusions of the 

1996 and 2006 CDs that ambient O3 concentrations can reduce the yield of major commodity 

crops in the U.S. and to support the use of crop E-R functions based on OTC experiments. 

Agricultural areas that would be likely to have the most significance to the public welfare would 

be those high production areas for sensitive crops that also are exposed to high O3

concentrations, such as areas in the Midwest and Mississippi Valley regions. 

5.3.2 Exposure/Risk-based Considerations 

Two main analyses are conducted in the WREA to estimate O3 impacts related to crop 

yield. Annual yield losses are estimated for 10 commodity crops and these estimates are then 

additionally used to estimate O3 impacts on producer and consumer economic surpluses (Table 

5-6 below; U.S. EPA, 2014a, sections 6.2, 6.5). 

Table 5-6. Exposure, risk and ecosystem services analyses related to crop yield. 

Crop-level impactB Agri-Ecosystem ServicesC

WREA estimatesA Annual Relative Yield Loss for 

Corn, Cotton, Potato, Sorghum, 

Soybean, Winter Wheat 

Economic surplus to crop producers 

and consumers 

A See Chapter 4 WREA; B See Section 6.2 WREA; C See Sections 6.4 and 6.5 WREA 
5-49 

 For what air quality scenarios were exposures and risks estimated? What 

approaches were used to estimate W126 exposures for those conditions? What 

are associated limitations and uncertainties? 

The WREA crop analyses described here were performed for five air quality scenarios 

using the methodology summarized in Table 5-4 above. In general, this methodology is identical 

to the air quality scenarios for the biomass loss analyses and have the same uncertainties and 

limitations summarized in section 5.2.2 above. These air quality scenarios described in more 

detail in the WREA (U.S. EPA, 2014a, chapter 4 and Appendix 4A). 

 What is the nature and magnitude of the cumulative exposure- and risk-related 

estimates for crop yield loss associated with remaining upon simulating just 

meeting the current O3 standard? What are the uncertainties associated this 

information? 

The WREA presents estimates of crop yield loss for the five air quality scenarios 

described above using 10 robust E-R functions for commodity crops that are grown across the 

U.S. (U.S. EPA, 2014a, section 6.5). The largest reduction in O3 induced crop yield loss occurs 

when moving from the recent conditions scenario to that for just meeting the current standard 

(U.S. EPA, 2014a, section 6.5). In the analyses for agricultural harvest, the largest estimates of 

yield changes also occur when comparing the recent conditions scenario to that for the current 

standard. Under recent conditions, the West, Southwest, and Northeast regions generally have 

the highest yield losses. For the 3-year average W126 scenarios, relative yield losses for winter 

wheat19 are less than one percent. For soybeans, yield losses for these scenarios range from just 

above 1 percent to below one percent (U.S. EPA 2014a, section 6.5). However, when evaluated 

at the county level, 99% of soybean producing counties (1,718) have greater than 5% yield loss 

under recent conditions, while no counties show yield loss at or above this level when air quality 

is adjusted to just meet the current standard (U.S. EPA, 2014a, section 6.5). 

The WREA estimates of O3-attributable percent yield loss based on 3-year average W126 

index values estimated after just meeting the current standard are relatively small (0.0 – 2.72%) 

across the 10 crop species analyzed, U.S. EPA 2014a, section 6.5, Appendix 6B). In considering 

these estimates, we recognize the significant uncertainties associated with several aspects of the 

analyses. Because the W126 estimates generated in the air quality analyses are inputs to the 

vegetation risk analyses for crop yield loss, any uncertainties in the air quality analyses are 

propagated into the those analyses (U.S. EPA 2014a, Table 6-27, section 8.5). 

 

19 Among the major crops, because winter wheat and soybeans are more sensitive to ambient O3 levels than 

other crops we focus on these crops for this discussion. 
5-50 

 To what extent are the exposures and risks remaining upon simulating just 

meeting the current O3 standard important from a public welfare perspective? 

From a public welfare prospective, the O3 attributable risks to crops estimated for 

conditions that just meet the current standard are small. However, it is unclear how much weight 

to put on these results given the multiple areas of uncertainty associated with these estimates as 

discussed in the WREA and summarized above, including those associated with the model-based 

adjustment methodology and those associated with projection of yield loss at the estimated O3

concentrations (U.S. EPA, 2014a, Table 6-27, section 8.5). In addition we note that while having 

sufficient crop yields is of high public welfare value, important commodity crops are typically 

heavily managed to produce optimum yields. Given all of the inputs that go into achieving these 

yields, such as fertilizer, herbicides, pesticides, and irrigation, it difficult to determine at what 

point O3-induced yield loss creates an adverse impact for the producer in the way of requiring 

increased inputs in order to maintain the desired yields. In contrast, based on the economic 

theory of supply and demand, increases in crop yields would be expected to result in lower prices 

for affected crops and their associated goods, which would benefit consumers. However, due to 

pre-existing market forces and subsidies, it is not clear that such benefits would be realized by 

the consumer. Given these competing impacts on producers and consumers, it is unclear how to 

determine what type of effect may be adverse to the public welfare. In considering this issue, 

CASAC states that “calculation of consumer and producer surpluses is a useful contribution to 

quantification of welfare effects. However, this national-level approach does not adequately 

account for negative effects on individual farmers and forest owners in high-ozone areas...” 

(Frey, 2014, p. 10). Instead, CASAC states that “[a] county scale is appropriate for assessing 

crop yield loss. Calculating producer and consumer surpluses at national or large region scales 

does not provide adequate protection. Farmers growing sensitive crops in high ozone locations 

can be considered a ‘sensitive population’ for welfare impacts, and crop yields under these 

conditions should be protected.” (Frey, 2014, pp. 14 – 15). The final WREA includes a countylevel analysis in Appendix 6B finding that 99 percent of soybean-producing counties, for 

example, exceed 5% yield loss under recent conditions, while no counties have relative yield 

losses above 5% for any crop after adjusting air quality scenarios to just meet the current 

standard. 

 What are the ecosystem services potentially affected by O3-related crop yield loss 

and to what extent are they important from a public welfare perspective? To 

what degree can the magnitude of the O3 effect on these services be qualitatively 

or quantitatively characterized? 

The WREA presents national-scale estimates of the effects of biomass loss on timber 

production and agricultural harvesting, which supply provisioning services of food and fiber, as 
5-51 

well as on carbon sequestration (U.S. EPA 2014a, section 6.5). Because the forestry and 

agriculture sectors are related and trade-offs occur between the sectors, the WREA also 

calculated the resulting market-based welfare effects of O3 exposure in the forestry and 

agriculture sectors. Overall effect on agricultural yields and producer and consumer surplus 

depends on the (1) ability of producers/farmers to substitute other crops that are less O3 sensitive, 

and (2) responsiveness, or elasticity, of demand and supply (U.S. EPA, 2014a, sections 6.5, 

8.2.1.3). Estimated O3-attributable economic welfare impacts on agricultural sectors associated 

with air quality conditions adjusted to just meet the existing and potential alternative W126 

standard levels were largely similar between the forestry and agricultural sectors. Estimates of 

consumer surplus, or consumer gains, were generally higher under those conditions (compared to 

recent conditions) in both sectors because higher productivity under lower O3 concentrations 

increased total yields and reduced market prices (U.S. EPA 2014a, Table 6-18). Because 

demand for most forestry and agricultural commodities is not highly responsive to changes in 

price, generally producer surplus, or producer gains, decline. For agricultural welfare, annualized 

combined consumer and producer surplus gains were estimated to be $2.6 trillion for modelbased adjustment to meet the current standard. Combined gains were essentially unchanged in 

comparisons of the current standard scenario to the average W126 scenario for 15 ppm-hrs, but 

gains increased by $21 million for the W126 scenario for 11 ppm-hrs and $231 million for the 

W126 scenario for 7 ppm-hrs. In some cases, lower prices reduce producer gains more than can 

be offset by higher yields (U.S. EPA, 2014a, Table 6-18). 

The WREA discusses multiple areas of uncertainty associated with these estimates (also 

summarized above), including those associated with the model-based adjustment methodology as 

well as those associated with projection of yield loss at the estimated O3 concentrations (U.S. 

EPA, 2014a, Table 6-27, section 8.5). 

5.4 VISIBLE FOLIAR INJURY 

Visible foliar injury resulting from exposure to O3 has been well characterized and 

documented over several decades of research on many tree, shrub, herbaceous, and crop species 

(U.S. EPA, 2013, 2006, 1996, 1984, 1978). The significance of O3 injury at the leaf and whole 

plant levels depends on an array of factors, and therefore, it may be difficult to quantitatively 

relate visible foliar injury symptoms to other vegetation effects such as individual tree growth, or 

effects at population or ecosystem levels (U.S. EPA, 2013, p. 9-39). Visible foliar injury by 

itself, however, is an indication of phytotoxicity due to O3 exposure and can impact the public 

welfare through damaging or impairing the intended use of the affected entity or the service it 

provides. For example, ways by which O3-induced visible foliar injury may impact the public 

welfare include: 1) visible damage to ornamental species used in landscaping or leafy crops 
5-52 

(spinach, lettuce, tobacco) that affects the economic value, yield, or usability of that plant (U.S. 

EPA 2007, section 7.4.1; Abt Associates, Inc., 1995); 2) visible damage to plants with special 

cultural significance (e.g., those used in tribal practices); 3) visible damage to species occurring 

in natural settings valued for their scenic beauty and/or recreational appeal, including in areas 

specially designated for more protection (e.g., federal Class I areas) (73 FR 16490). Given 

limitations in the available information pertaining to the first two categories, 20 the discussions of 

the evidence and exposure/risk information in sections 5.4.1 and 5.4.2 below focus primarily on 

what is known about visible foliar injury that has been shown to occur in natural settings valued 

for their scenic beauty and/or recreational appeal. 

At the time of the last review, the following was known: 

1) Ozone causes diagnostic visible injury symptoms on studied bioindicator species. 

2) Soil moisture is a major confounding effect that can decrease the incidence and 

severity of visible foliar injury under dry conditions and vice versa. 

3) The most extensive dataset regarding visible foliar injury incidence across the U.S. 

was that collected by the USFSFHM/FIA Program. 

4) Visible foliar injury incidence was considered to be widespread in both the eastern and 

western U.S. based on staff analyses of county level air quality data and USFS 

biomonitoring data which showed that for each year within a four year period (2001 – 

2004) the percentage of counties having a biosite with visible foliar injury ranged 

between 11-30% at an 8-hour average annual level of 0.074 ppm (U.S. EPA, 2007, 

section 7.6.3.2). 

 In the remainder of this section, we consider how the currently available evidence and 

exposure/risk information informs our understanding of the relationship that exists between 

visible foliar injury and exposures to O3 in ambient air and consideration of the adequacy of 

protection provided by the current standard. The policy-relevant evidence and weight-ofevidence conclusions drawn from the ISA are discussed in section 5.4.1, and the exposure/risk 

and associated ecosystem services estimates from the WREA, are discussed in section 5.4.2. 

Important uncertainties and limitations in each type of available information are also discussed in 

these two sections. 

 

20 Qualitative information regarding potential cultural impacts of O3-induced visible foliar injury is noted in 

section 5.5 and Appendix 5-A). 
5-53 

5.4.1 Evidence-based Considerations 

 To what extent has scientific information become available that alters or 

substantiates our previous conclusions of O3-related visible foliar injury and of 

factors that influence associations between O3 exposures or concentrations and 

visible foliar injury? 

Recent research continues to build and substantiate the previous conclusions and findings 

drawn from several decades of research on many tree, shrub, herbaceous, and crop species (U.S. 

EPA, 2013, 2006, 1996, 1984, 1978) that O3-induced visible foliar injury symptoms are well 

characterized and considered diagnostic on certain bioindicator plant species. Diagnostic usage 

for these plants has been verified experimentally in exposure-response studies, using exposure 

methodologies such as continuous stirred tank reactors (CSTRs), open-top chambers (OTCs), 

and free-air fumigation (FACE). Although there remains a lack of robust exposure-response 

functions that would allow prediction of visible foliar injury severity and incidence under 

varying air quality and environmental conditions, experimental and observational evidence has 

clearly established a consistent association of the presence of visible injury symptoms with O3

exposure, with greater exposure often resulting in greater and more prevalent injury (U.S. EPA 

2013, section 9.4.2). This new research includes: 1) controlled exposure studies conducted to 

test and verify the O3 sensitivity and response of potential new bioindicator plant species; 2) 

multi-year field surveys in several National Wildlife Refuges (NWR) documenting the presence 

of foliar injury in valued areas; 3) ongoing data collection and assessment by the USFS 

FHM/FIA program, including multi-year trend analysis (U.S. EPA 2013, section 9.4.2). These 

recent studies, in combination with the entire body of available evidence, thus form the basis for 

the ISA determinations of a causal relationship between ambient O3 exposure and the occurrence 

of O3-induced visible foliar injury on sensitive vegetation across the U.S. (U.S. EPA 2013, p. 9-

42). 

With regard to evidence from controlled exposure studies, a recent study of 28 plant 

species confirmed prior findings of O3 causing predictable diagnostic visible foliar injury 

symptoms on some species of plants. This study selected 28 plant species, most of which grow 

naturally throughout the northeast and midwest US, including in national parks and wilderness 

areas, that were suspected of being O3-sensitive, and exposed them to four different O3

concentrations (30, 60, 90, and 120 ppb) in CSTR chambers (Kline et al., 2008). Two 

experiments were conducted in each year of the study (2003 and 2004). Plants were exposed to 

O3 for 7 hours a day, five days a week over the course of each experiment. Specifically, in 2003, 

the first experiment lasted from July 14 to August 21 and included 29 days of O3 exposure and 

the second from September 9 to 30 and included 16 exposure days. In 2004, the first experiment 

was conducted from July 13 to August 10 with 21 O3 exposure days and the second from August 
5-54 

27 to September 24, including 21 days of O3 exposure. Though the exposures were cumulative 

over the course of the study, exposures were reported only in terms of the target exposure 

concentration for each experiment. The study reported O3-induced responses in 12, 20, 28 and 

28 of the 28 tested species at the 30, 60, 90 and 120 ppb exposure concentrations21, respectively. 

Based on their findings, the authors suggest that American sycamore, aromatic sumac, bee balm, 

buttonbush, common milkweed, European dwarf elderberry, New England aster, snowberry and 

swamp milkweed would make the most useful bioindicator species. Some of these species are 

native to Class 1 areas (discussed further below). The staff additionally concludes that given that 

the exposure protocol was designed to create a continuous exposure level, not a fluctuating one, 

this study shows that O3-induced foliar injury can occur from 7-hour exposures repeated over 

multiple days at O3 concentrations that are below the 75 ppb level of the current standard.22 

While this type of controlled study provides clear evidence of cause and effect, it also has 

limitations. The authors, recognizing this cautioned that “extrapolation of these CSTR results to 

the field must be done carefully, since CSTR/greenhouse conditions ... are not representative of 

natural environmental conditions” (Kline et al., 2008). 

A string of recently published multi-year field studies provide a complimentary line of 

field-based evidence by documenting the incidence of visible foliar injury symptoms on a variety 

of O3-sensitive species over multiple years and across a range of cumulative, seasonal exposure 

values in several eastern and midwestern NWRs (U.S. EPA 2013, section 9.4.2.1; Davis and 

Orendovici, 2006; Davis, 2007a, b; Davis, 2009). Some of these studies also included 

information regarding soil moisture stress using the Palmer Drought Severity Index (PDSI). 

While environmental conditions and species varied across the four NWRs, visible foliar injury 

was documented to a greater or lesser degree at each site. As discussed further below, visible 

foliar injury incidence in these types of areas has greater significance to the public welfare. 

 To what extent have important uncertainties identified in the last review been 

reduced and/or new uncertainties emerged? 

The studies mentioned above also provide additional information regarding an important 

uncertainty identified in the previous review, i.e., the role of soil moisture in influencing visible 

foliar injury response (U.S. EPA 2013, section 9.4.2). These studies confirm that adequate soil 

moisture creates an environment conducive to greater visible foliar injury in the presence of O3

 

21 Two of the target exposure levels, 30 and 60 ppb, fall below the level of the current standard (75 ppb). 

The mean exposure concentrations achieved in the CTSRs for the 30 ppb target level for each year and study were 

27.9, 26.3, 27.1, and 29.3 ppb and for the 60 ppb target level were 56.6, 55.8, 57.9, and 59.0 ppb, for 2003 study 1, 

2003 study 2, 2004 study 1, and 2004 study 2, respectively. 

22 The current standard is met when the 3-year average of the 4th highest daily maximum 8-hour average 

concentrations is at or below 75 ppb. 
5-55 

than drier conditions. As stated in the ISA, “[a] major modifying factor for O3-induced visible 

foliar injury is the amount of soil moisture available to a plant during the year that the visible 

foliar injury is being assessed ... because lack of soil moisture generally decreases stomatal 

conductance of plants and, therefore, limits the amount of O3 entering the leaf that can cause 

injury” (U.S. EPA, 2013, p. 9-39). As a result, “many studies have shown that dry periods in 

local areas tend to decrease the incidence and severity of O3-induced visible foliar injury; 

therefore, the incidence of visible foliar injury is not always higher in years and areas with higher 

O3, especially with co-occurring drought (Smith, 2012; Smith et al., 2003)” (U.S. EPA, 2013, p. 

9-39). This “...partial ‘protection’ against the effects of O3 afforded by drought has been 

observed in field experiments (Low et al., 2006) and modeled in computer simulations 

(Broadmeadow and Jackson, 2000)” (U.S. EPA, 2013, p. 9-87). In considering the extent of any 

protective role of drought conditions, however, the ISA also notes that other studies have shown 

that “drought may exacerbate the effects of O3 on plants (Pollastrini et al., 2010; Grulke et al., 

2003)” and that “[t]here is also some evidence that O3 can predispose plants to drought stress 

(Maier-Maercker, 1998)”. Accordingly, the ISA concludes that “the nature of the response is 

largely species-specific and will depend to some extent upon the sequence in which the stressors 

occur” (U.S. EPA, 2013, p. 9-87). Such uncertainties associated with describing the potential for 

foliar injury and its severity or extent of occurrence for any given air quality scenario due to 

confounding by soil moisture levels make it difficult to identify an appropriate degree of annual 

protection (as well as ambient O3 exposure conditions that might be expected to provide that 

protection). 

 To what extent does the available evidence indicate the occurrence of O3-related 

visible foliar injury attributable to cumulative exposures at lower ambient O3

concentrations than previously established or to exposures at or below the level of 

the current standard? 

Recently available evidence confirms the evidence available in previous reviews that 

visible foliar injury can occur when sensitive plants are exposed to elevated O3 concentrations in 

a predisposing environment (i.e., adequate soil moisture (U.S. EPA, 2013, section 9.4.2). Recent 

evidence also continues to indicate the occurrence of visible foliar injury at cumulative ambient 

O3 exposures previously established. Since the 2006 O3 CD, results from several multi-year field 

surveys and experimental screenings of O3-induced visible foliar injury on vegetation also show 

that visible foliar injury can occur under conditions where the annual 8-hour average O3

concentrations are at or below the level of the current standard, as discussed here. Limited 

information exists regarding the incidence of visible foliar injury occurring in areas that have 

design values that meet the current 3-year average 8-hour standard. 
5-56 

To facilitate comparison with other studies reporting foliar injury response to W126 

cumulative exposures, we obtained air quality data from the EPA’s AQS database for monitors 

in each study location and calculated the 12-hr W126 index values and obtained the maximum 

4th highest 8-hour average values for a subset of the most recent years included in each study 

(Table 5-7). As the shaded rows in Table 5-7 below show, in the years 2002/2003 and 2004 in 

the Cape Romain NWR in South Carolina, and the Seney NWR in Michigan, respectively, the 4th

highest daily maximum 8-hour average O3 concentrations were at or below the level of the 

current standard. We additionally note that the Cape Romain site met the current standard of 75 

in every 3-year period during the study and has consistently met the standard from 2001 to 

2012.23 Under these air quality conditions, three species (i.e., winged sumac, Chinese tallow 

tree, and wild grape) exhibited O3-induced stipple. In 2002, 32% of the examined wild grape 

plants, 20% of the winged sumac plants, and 4.6% of the Chinese tallow tree plants, respectively, 

were symptomatic (Davis, 2009). At the same time, the 12-hour W126 index value was 20 ppmhrs. In 2003, when air quality was somewhat improved, foliar injury declined, with only 13.3% 

of wild grape showing O3 stipple at a maximum 4th highest 8-hour of 74 ppb and a W126 index 

value of 11 ppm-hrs. The PDSI values were 0.27 and 2.45 in 2002 and 2003, respectively. 

These values show that 2003 was a wetter year than 2002, though 2002 would have been 

considered within the normal soil moisture range. 

At the Seney NWR site, by comparison, the annual W126 level was similar in 2004 to 

that at Cape Romain in 2003, and the annual 8-hour average level was below that of the current 

standard, though the 3-year average design values were above that of the current standard for that 

year. Not surprisingly, given the lower O3 air quality in 2004, the Seney study reported injury 

ranging from about 2% on common milkweed to about 6% on spreading dogbane. Though this 

study does not provide the PDSI values, the authors provided some discussion of a possible 

relationship stating that “the incidence of ozone injury on spreading dogbane, but not other 

species, was weakly, but not significantly, related to the drought index (PDSI)....However this 

relationship was too weak to be used for predictive purposes” (Davis, 2007b). The authors then 

conclude that “[n]evertheless, the threshold SUM06 ozone level needed to induce stipple on 

sensitive plants within the Seney refuge is likely 5000 ppb-hrs under the environmental 

conditions of these surveys” (Davis, 2007b). On the basis of the above, the staff concludes that 

these studies confirm that visible foliar injury has been shown to occur in the field at W126 

index values ranging down to 10 ppm-hrs and provide limited evidence that such foliar injury 

 

23 Design values (concentrations in the form of the standard) for this monitoring site during this period are 

presented in the file available at: http://www.epa.gov/airtrends/values.html (US EPA, 2014d). 
5-57 

can occur in areas with special public welfare significance during periods that meet the current 

standard. 

Table 5-7. Visible foliar injury incidence in four National Wildlife Refuges. 

Name/ Site #/ Ref.A Year B 4th highest daily maximum 

8-hour average 

12-hour 

W126 

% Plants with 

visible injury 

Cape Romain NWR, South Carolina / 

450190046 (Davis, 2009) 

2002 0.075 ppm 20 ppm-hrs 5 - 32 

2003 0.074 ppm 11 ppm-hrs 3 - 13 

Moosehorn NWR, Maine/ 230090102 

(Davis, 2007a) 

2002 0.1 ppm 24 ppm-hrs 0 - 17 

2003 0.083 ppm 22 ppm-hrs 0 - 13 

2004 0.082 ppm 14 ppm-hrs 3 - 10 

Seney NWR, Michigan/ 261530001 

(Davis, 2007b) 

2002 0.083 ppm 11 ppm-hrs 0 - 13 

2003 0.076 ppm 15 ppm-hrs 1 - 6 

2004 0.074 ppm 10 ppm-hrs 2 - 6 

Brigantine NWR, New Jersey / 

340010005/ 

(Davis and Orendovici, 2006) 

2001 0.095 ppm 39 ppm-hrs 0 – 45 

2002 0.092 ppm 53 ppm-hrs 0 – 4 

2003 0.085 ppm 36 ppm-hrs 0 - 4 

A Studies (cited above) reported exposures in terms of SUM06 form. EPA staff, using AQS data for the same 

monitors, calculated exposures in terms of the current 8-hour and W126 forms: 

http://www.epa.gov/ttn/airs/airsaqs/ (US EPA, 2014b) B Only recent years with available W126 data were included in the table. 

By far the most extensive field-based dataset of visible foliar injury incidence is that 

obtained by of the USFS FHM/FIA biomonitoring network program. A trend analysis of data 

from the sites located in the Northeast and North Central U.S. for the 16 year period (1994-2009) 

(Smith, 2012) provides additional evidence of foliar injury occurrence in the field as well as 

some insight into the influence of changes in air quality and soil moisture on visible foliar injury 

and the difficulty inherent in predicting foliar injury response under different air quality/soil 

moisture scenarios (Smith, 2012; U.S. EPA 2013, section 9.2.4.1). In this study ambient 

exposures were expressed in terms of the SUM06 cumulative index coupled with a measure of 

the number of peak hourly concentrations above 100 ppb (N100). Soil moisture estimates were 

generated using both the PDSI and the plant moisture availability index (MI). Foliar injury was 

expressed in terms of the biosite index (BI)24. The authors observed that over this 16-year 

 

24 Biosite index (BI) is the average score (proportion of leaves with injury [“amount”] x mean severity of 

symptoms on injured leaves [“severity”]) for each species averaged across all species on the biosite multiplied by 

1,000. 
5-58 

period, “injury indices have fluctuated annually in response to seasonal ozone concentrations and 

site moisture conditions. Sites with and without injury occur at all ozone exposures but when 

ambient concentrations are relatively low, the percentage of uninjured sites is much greater than 

the percentage of injured sites; and regardless of ozone exposure, when drought conditions 

prevail, the percentage of uninjured sites is much greater than the percentage of injured sites” 

(Smith, 2012). The authors further note that while “both site moisture and ozone exposure play a 

role in foliar injury expression ... the interplay among these three factors is unique for each year 

and possibly each site. Extreme moisture deficits decrease foliar injury, ... [and] ... [i]n no year 

do high ozone exposures override the controlling effect of site moisture, although at the other 

end of the scale, injury severity is minimized under conditions of low ozone exposure regardless 

of site moisture conditions. This implies a necessary threshold of ozone exposure for injury to 

occur....” “In a similar analysis, Rose and Coulston (2009) reported a high percentage of 

biosites with injury across the Southern region in 2003, a year when SUM06 values >10 ppm-h 

were widespread at the same time that the land area was in moisture surplus or balance.” Thus, 

Rose and Coulston (2009) also “found evidence that it is the co-occurrence of sufficient moisture 

and elevated ozone that determine whether injury occurs to bioindicator plants, not ozone 

exposure alone.” Regarding the role of peak O3 concentrations (>100 ppb O3), Smith (2012) 

reported that over the 16-year period concentrations above 100 ppb have declined, and that this 

“... may account for the observed decrease in the severity of ozone-induced foliar injury to ozone 

sensitive bioindicator plants in eastern forests.” They also note that “[t]here is no compelling 

evidence, however, that moderate ozone concentrations, as reflected in seasonal mean SUM06 

data, are on the decline” and “[t]his may explain why injury continues to be detected on many of 

the same sites every year” (Smith, 2012). The authors thus conclude that, “[a]lthough it is 

reasonable to remain concerned about long-term impacts of ozone pollution on our forest 

ecosystems, the findings of this biomonitoring survey point to a declining risk of probable 

impact on eastern forests over the 16-year period from 1994 to 2009” (Smith, 2012). 

In a similar assessment of the USFS FHM/FIA data in the West, six years (2000 to 2005) 

of biomonitoring data for O3 injury were evaluated for the three coastal states of California, 

Oregon and Washington (Campbell et al., 2007; U.S. EPA 2013, section 9.4.2.1). Campbell et 

al., 2007 found that “...ozone injury occurs frequently (25 to 37 percent of sampled biosites) in 

California forested ecosystems demonstrating that ozone is present at phytotoxic levels.” This 

study concluded that, “in California, an estimated 1.3 million acres of forest land and 596 million 

cubic feet of wood are at moderate to high risk to impacts from ozone. However, [m]ore years of 

data are needed to discern any trends” (Campbell et al, 2007). Though this study does not 

discuss the role of soil moisture in describing the results, the criteria used to select the 

biomonitoring sites include one that considers soil conditions. The best sites are identified as 
5-59 

those with low drought potential and good fertility. Thus, given the relatively high O3

concentrations that occur in California and the likelihood that many of the biomonitoring sites 

occur in areas that have sufficient soil moisture, the high percentage of sampled biosites with 

foliar injury is not unexpected.25 

These recent studies continue to provide evidence of O3-induced foliar injury occurring in 

many areas across the U.S. and augment our understanding of O3-related visible foliar injury and 

of factors that influence associations between O3 exposures or concentrations and visible foliar 

injury such as soil moisture. 

 To what extent does currently available evidence suggest locations where the 

vulnerability of sensitive species, ecosystems and/or their associated services to 

O3-related visible foliar injury would have special significance to the public 

welfare? 

As mentioned above, federally designated Class I areas are afforded stringent protections 

under the 1977 amendments to the CAA. The CAA gives federal land managers of Class I areas 

“the responsibility to protect all air quality related values (AQRVs)...from deterioration.... In 

order to determine if deterioration is occurring, baseline AQRVs must be established” (Davis, 

2009). Because of this need and the significance of these areas, studies often focus on these 

sites. For example, a study by Kohut (2007) was undertaken to assess the risks of O3-induced 

visible foliar injury on O3-sensitive vegetation in 244 parks managed by the NPS (U.S. EPA, 

2013, pp. 9-40 to 9-41, U.S. EPA, 2014a, pp. 7-19 to 7-20). Kohut (2007) concluded that the 

risk of visible foliar injury was high in 65 parks (27 percent), moderate in 46 parks (19 percent), 

and low in 131 parks (54 percent). Thus, while this study suggests that there may be a reason for 

concern in as much as 46% of the parks, there were a number of important limitations associated 

with this study (described in footnotes 8 and 9 below) that weakened this conclusion. Given the 

importance of this kind of assessment, the WREA used Kohut (2007) as the conceptual basis for 

the subsequent WREA screening-level assessment, though numerous modifications were made 

to the approach to make it applicable to the context of this O3 NAAQS review (see section 5.4.2 

below). 

In addition, as described above, several recently published studies (U.S. EPA 2013, 

section 9.4.2.1; Davis and Orendovici, 2006; Davis, 2007a,b; Davis, 2009, Kohut, 2007) were 

conducted in federally protected areas including federally designated Class I areas such as 

national parks. These studies confirm that visible foliar injury has been observed in these areas 

under annual air quality conditions with ambient concentrations at or below the level of the 

 

25 Staff additionally notes that a large proportion of O3 monitoring sites in California did not meet the 

current standard during the study period (see: http://www.epa.gov/airtrends/values.html) (US EPA, 2014d). 
5-60 

current standard and at W126 index values within the CASAC range recommended in past 

reviews. This evidence continues to suggest that O3-sensitive species and their associated 

ecosystems and services continue to remain vulnerable to visible foliar injury incidence in areas 

that have been afforded special protection by Congress and that have special significance to the 

public welfare. 

5.4.2 Exposure- and Risk-based Considerations 

The WREA presents a number of analyses considering air quality conditions associated 

with increased prevalence of visible foliar injury and potential associated welfare impacts (see 

Table 5-8 below, U.S. EPA, 2014a, Chapter 7). An initial analysis included the development of 

benchmark criteria reflecting different prevalence of visible foliar injury in conjunction with 

different W126 exposures and in some cases, soil moisture conditions. These criteria were then 

used in a screening-level assessment to characterize potential risk of foliar injury incidence under 

2006-2010 conditions in 214 national parks. The last analysis was a case study assessment on 

three national parks, which also provides limited characterization of the associated ecosystem 

services. Despite the limitations and uncertainties associated with these analyses, and 

recognizing that the recent air quality conditions in most cases (prior to any model-based 

adjustment) did not meet the current standard, we believe that they help inform our 

understanding of the relationship between soil moisture and foliar injury incidence, as well as 

provide limited support for our conclusions regarding risk of visible foliar injury incidence under 

air quality conditions likely to meet the current standard in areas of special significance to the 

public welfare. 
5-61 

Table 5-8. Exposure, risk and ecosystem services analyses related to visible foliar 

injury. 

Ecosystem Level Effects Ecosystem Services 

WREA 

estimates 

Proportion of FHM/FIA biosites with 

foliar injury incidence at various 

W126 index values and soil moisture 

levels 

Percent of 214 national parks 

exceeding various W126 benchmarks 

derived from FHM/FIA biosite 

analysis A 

Case study of 3 national parks characterized impacts 

using available visitor and use data, including monetary 

data for activities and visitor expenditures: 

 Utilized Willingness-to-Pay studies for scenic 

impairment; 

 Assessed the overall cover of sensitive species; 

 Compared sensitive species cover to trails and 

overlooks; and, 

 Estimated percent of park area with O3 concentrations 

above different W126 index values averaged over 

three consecutive years. 

A The screening-level assessment of 214 national parks additionally included observations based on the modelbased adjustments to just meet the current standard and targets for the three W126 scenarios (discussed below) 

but did not conduct a full analysis using these data. 

 For what air quality scenarios were exposures and risks estimated? What 

approaches were used to estimate W126 exposures for those conditions? What 

are associated limitations and uncertainties? 

Three types of foliar injury analyses were performed in the WREA and are considered 

below. They include an analysis using USFS FHM/FIA biosite data, a screening-level 

assessment in 214 national parks, and case studies of three national parks. The analysis of USFS 

biosite data was done using O3 concentrations estimated for a national-scale surface of 

concentrations (at a 12 x 12 km2

 grid cell resolution in contiguous U.S.) using interpolation 

methodology applied to concentrations at O3 monitor locations (U.S. EPA, 2014a, section 4.3.2, 

Appendix 4A). The analysis of USFS FHM/FIA data and the screening-level analysis using 

W126 benchmarks derived from these data used surfaces for each year from 2006 through 2010 

(U.S. EPA, 2014a, Appendix 4A, section 4.2). In the National Park case study analyses, 

observations related to air quality were made for five air quality scenarios by the methodology 

summarized in Table 5-4 above.26 

The W126 index values in the individual years from 2006 to 2010 at monitors ranged 

from less than 5 ppm-hrs up to above 48 ppm-hrs (U.S. EPA, 2014a, Figure 4-4 and Table 4-3). 

 

26 In general, this methodology involved two steps. The first is derivation of the average W126 value 

(across the three years) at each monitor location. This value is based on unadjusted data for recent conditions and 

adjusted concentrations for the 4 other scenarios. The development of adjusted concentrations was done for each of 

9 regions independently (see U.S. EPA, 2014a, section 4.3.4.1). In the second step, national-scale spatial surfaces 

(W126 values for each model grid cell) were created using the monitor-location values and the VNA spatial 

interpolation technique (details on the VNA technique are presented in U.S. EPA, 2014a Appendix 4A). 
5-62 

Concentration estimates varied appreciably across the five years with the median index values 

across grid cells ranging from a low of 5.5 ppm-hrs in 2009 up to 11 ppm-hrs in 2006 (U.S. EPA, 

2014a, Appendix 4A, section 4.2). During the recent conditions period (2006 through 2008), the 

average W126 index values (across the three-year recent conditions period) at the monitor 

locations ranged from below 10 ppm-hrs to 48.6 ppm-hrs (U.S. EPA 2014a, Figure 4-4 and 

Table 4-3). After adjusting the 2006-2008 data to just meet the current standard in each region, 

and subsequent application of the VNA technique to the current standard scenario monitor 

location values, the average W126 index values were below 15 ppm-hrs across the national 

surface with the exception of a very small area of the southwest region (near Phoenix) where the 

average W126 index values was near or just above 15 ppm-hrs. A lowering of the highest values 

occurred with application of the interpolation method as a result of estimating W126 index 

values at a 12 x 12 km2

 grid resolution rather that at the exact location of a monitor. This 

indicates one uncertainty associated with this aspect of the approach to estimating W126 index 

values for the adjusted air quality just meeting the current standard. Other uncertainties are 

summarized in section 5.2.2 above. 

 What are the nature and magnitude of the cumulative exposure- and risk-related 

estimates for visible foliar injury under recent conditions or conditions meeting 

the current O3 standard? 

As an initial matter, we consider the analysis of the biomonitoring site data from the 

USFS FHM/FIA Network, described in section 7.2 of the WREA.27 Using this dataset and 

associated data for soil moisture during the sample years along with ambient air O3

concentrations based on monitoring data from 2006 to 2010 and spatial interpolation 

methodology (as described above), the proportion of biosites with any foliar injury are observed 

to increase with increasing annual W126 index values up to specific values after which there is 

little change in proportion of affected biosites with higher W126 index values (see Figure 5-5 

below; U.S. EPA, 2014a, section 7.2, Figure 7-10). The proportion of biosites metric is derived 

by first ordering the data (across biosites and sample years) by W126 index value estimated for 

that biosite and year. Then for each W126 index value the proportion of biosites exceeding the 

selected biosite index value for all observations at or below that W126 index value is calculated. 

The WREA repeated this using a biosite index value greater than zero, indicating presence of any 

foliar injury (USFS, 2011). 

When looking only at presence or absence of foliar injury (“any injury”) with the 

exception of 2008, the proportion of biosites across all W126 index values with foliar injury 

 

27 Data were not available for several western states (Montana, Idaho, Wyoming, Nevada, Utah, Colorado, 

Arizona, New Mexico, Oklahoma, and portions of Texas). 
5-63 

exceeds 15 percent; in 2006, it exceeds 20 percent, while in 2008 the proportion of biosites with 

foliar injury across all W126 index values was just below 15 percent (U.S. EPA, 2014a, section 

7.2.3, Figure 7-9). When categorized by moisture levels, the data demonstrate a distinct pattern. 

In general, the WREA concludes that the results of these foliar injury analyses demonstrate a 

similar pattern –the proportion of biosites showing the presence of any foliar injury (biosite 

index >0) increases from zero to about 20% (Figure 5-5 below). This increase occurs with 

increasing W126 index values up to approximately 10 ppm-hrs for any foliar injury (biosite 

index >0), with little change in proportion of biosites with any injury at higher W126 index 

values. The data for biosites during normal moisture years are very similar to the dataset as a 

whole, with an overall proportion of close to 18 percent for presence of any foliar injury. Among 

the biosites with a relatively wet season (average Palmer Z => 1), the proportion of biosites 

showing injury is much higher and the relationship with annual W126 index value is much 

steeper. Much lower proportions of biosites are reached for the any injury category at biosites 

with relatively dry seasons (average Palmer Z < -1.24), potentially indicating that drought may 

provide some protection from foliar injury as discussed in the ISA (U.S. EPA, 2014a, section 

7.2.3, Figures 7-10). This information provides insight into the relationship between soil 

moisture and foliar injury and the issue of whether drought provides protection from foliar 

injury. Thus, there is relatively little change in the proportion of biosites beyond a W126 index 

value of 10 ppm-hrs. There are two important observations that can be made from these analyses: 

(1) the proportion of biosites exhibiting foliar injury rises rapidly at increasing W126 index 

values below approximately 10 ppm-hrs, and (2) there is relatively little change in the 

proportions above W126 index values of approximately 10 ppm-hrs. 
5-64 

Figure 5-5. Cumulative proportion of biosites with any foliar injury present, by moisture 

category (U.S. EPA 2014a, Figure 7-10). 

We additionally consider the WREA screening-level assessment in 214 parks in the 

contiguous U.S. that employed benchmark criteria developed from the above analysis (Table 5-

9).28,

29 For example, annual O3 concentrations corresponding to a W126 index value of 10.46 

ppm-hrs represents the O3 exposure concentration where the slope of exposure-response 

relationship changes for FHM biosites. The WREA refers to this as the “base scenario” 

benchmark. Above this index value, the percentage of FHM biosites showing foliar injury 

remains relatively constant. The W126 benchmarks across the five scenarios range from 3.05 

ppm-hrs (five percent of biosites, normal moisture, any injury) up to 24.61 ppm-hrs (10% of 

biosites, dry, any injury). For the scenario of 10% biosites with injury, W126 index values were 

approximately 4, 6, and 25 ppm-hours for wet, normal and dry years, respectively. The national-

 

28 The parks assessed here include lands managed by the NPS in the continental U.S., which includes 

National Parks, Monuments, Seashores, Scenic Rivers, Historic Parks, Battlefields, Reservations, Recreation Areas, 

Memorials, Parkways, Military Parks, Preserves, and Scenic Trails. 

29 The WREA applied different foliar injury benchmarks in this assessment after further investigation into 

the benchmarks applied in Kohut (2007), which were derived from biomass loss rather than visible foliar injury. 

Kohut cited a threshold of 5.9 ppm-hrs for highly sensitive species from Lefohn et al. (1997), which was based on 

the lowest W126 estimate corresponding to a 10% growth loss for black cherry. For soil moisture, Kohut (2007) 

qualitatively assessed whether there appeared to be an inverse relationship between soil moisture and high O3

exposure. 

0 10 20 30 40

0.00 0.05 0.10 0.15 0.20 0.25

Biosite Index > 0

W126 (ppm-hrs)

Proportion of sites

All

Wet

Normal

Dry
5-65 

scale screening-level assessment includes 42 parks with O3 monitors and a total of 214 parks 

with O3 exposure estimated from the interpolated O3 surface for individual years from 2006 to 

2010 (U.S. EPA, 2014a, Appendix 7A). These data were combined with lists from the NPS of 

the parks containing O3-sensitive vegetation species (NPS, 2003, 2006). Based on NPS lists, 95 

percent of the parks in this assessment contain at least one O3-sensitive species. This analysis for 

recent air quality conditions, estimates that 58 percent of parks exceeded the benchmark criteria 

corresponding to the base scenario (W126>10.46 ppm-hrs, 17.7 percent of biosites, all moisture 

categories, any injury) for at least three years in the period from 2006 to 2010 (U.S. EPA, 2014a, 

section 7.3.2). Based on model-based adjustments to meet the current standard, none of the 214 

parks have average W126 index values that would exceed the annual benchmark criteria for the 

base scenario (W126 >10.46 ppm-hrs) (U.S. EPA, 2014a, section 7.3.3.3). 

Table 5-9. Benchmark criteria for O3 exposure and relative soil moisture used in 

screening-level assessment of parks (from U.S. EPA 2014a, Table 7-6). 

Scenario Description 

W126 Benchmark (in ppm-hrs) 

Wet 

(Palmer Z ≥1) 

Normal Moisture 

(Palmer Z between 

-1.25 and 1) 

Dry 

(Palmer Z < -

1.25) 

Base 

17.7% of all FHM biosites showed any 

injury (higher W126 index values have a 

relatively constant percentage of FHM 

biosites showing injury) 

10.46 

(soil moisture not considered) 

5% of 

biosites 

5% of FHM biosites showed any injury, 

reflects soil moisture categorization 3.76 3.05 6.16 

10% of 

biosites 

10% of FHM biosites showed any injury, 

reflects soil moisture categorization 4.42 5.94 24.61 

15% of 

biosites 

15% of FHM biosites showed any injury, 

reflects soil moisture categorization 4.69 8.18 N/A 

20% of 

biosites 

20% of FHM biosites showed any injury, 

reflects soil moisture categorization 5.65 N/A N/A 

Lastly, we consider the WREA case study analysis that focused on characterizing the 

ecosystem services potentially associated with visible foliar injury in three specific national 

parks (case study assessment). The parks included were Great Smoky Mountains National Park 

(GRSM), Rocky Mountain National Park (ROMO), and Sequoia/Kings National Parks (SEKI). 

For each park, the potential impact of O3-related foliar injury on recreation (cultural services) 

was considered in light of information on visitation patterns, recreational activities and visitor 

expenditures. For example, visitor spending in 2011 exceeded $800 million, $170 million and 
5-66 

$97 million dollars in GRSM, ROMO and SEKI, respectively. This assessment also included 

percent cover of species sensitive to foliar injury and focused on the overlap between recreation 

areas within the park and elevated W126 concentrations. Ozone concentrations in GRSM have 

been among the highest in the eastern U.S. In the recent conditions scenario, the grid cells 

representing 44 percent of GRSM had three year average W126 index value above 15 ppm-hrs. 

After adjustments to just meet the current standard, no grid cell had a three-year average W126 

index value above 7 ppm-hrs. In the recent conditions scenario for ROMO, three-year average 

W126 index values for all grid cells were above 15 ppm-hrs. In the current standard scenario, 

values for 59 percent of the park were below 7 ppm-hrs. For SEKI, three-year average W126 

index values for all grid cells were above 15 ppm-hrs in the recent conditions scenario, but 

dropped below 7 ppm-hrs for the current standard scenario (U.S. EPA, 2014a, section 7.4). 

In summary, these analyses indicate that O3 concentrations in U.S. national parks in 

recent years correspond to W126 index values at which some foliar injury may occur, with 

variation associated with relative soil moisture conditions. None of the 214 parks assessed are 

estimated to exceed the annual benchmark criteria for the base scenario (W126 >10.46 ppm-hrs) 

after adjusting air quality to meet the current standard. Although adjusted scenarios to just meet 

the current standard indicate substantial reductions in three-year average W126 index values 

estimated by the VNA approach, some individual year values may range higher. The case study 

analysis of three parks indicates the potential for appreciable ecosystem services impact 

associated with foliar injury. While these analyses indicate the potential for foliar injury to occur 

under conditions that meet the current standard, the extent of foliar injury that might be expected 

under such conditions is unclear from these analyses. 

 To what extent are the exposures and risks remaining upon simulating just 

meeting the current O3 standard important from a public welfare perspective? 

The screening level assessment, as described above, indicates that risk of visible foliar 

injury is likely to be lower in most national parks after simulating just meeting the current 

standard. Based on the national-scale analysis, visible foliar injury would likely continue to 

occur at lower O3 exposures, including some sensitive species growing in areas (e.g., National 

Parks and other Class I areas) that may provide important cultural ecosystem services to the 

public. Staff notes that such occurrences might reasonably be considered to have some 

importance from a public welfare perspective, as discussed in section 5.1 above. 
5-67 

 What are the ecosystem services potentially affected by visible foliar injury, to 

what degree can the magnitude of these effects be qualitatively or quantitatively 

characterized, and to what extent are they important from a public welfare 

perspective? 

The ecosystem services most likely to be affected by O3-induced foliar injury are cultural 

services, including aesthetic value and outdoor recreation. Aesthetic value and outdoor 

recreation depend on the perceived scenic beauty of the environment. Many outdoor recreation 

activities directly depend on the scenic value of the area, in particular scenic viewing, wildlifewatching, hiking, and camping. These activities and services are of significant importance to 

public welfare as they are enjoyed by millions of Americans every year and generate millions of 

dollars in economic value (U.S. EPA, 2014a, Chapter 5, Chapter 7). These aesthetic values are at 

risk of impairment because of O3-induced damage directly due to foliar injury. Other ecosystem 

services that have also been found to be associated with O3-sensitive plants include those that fall 

under the categories of provisioning. For example, several tribes have indicated that many of the 

known confirmed O3-sensitive species (including bioindicator species) are culturally significant 

(see Appendix 5-A). Although data are not available to explicitly quantify these negative effects 

on ecosystem services, several qualitative analyses conducted in the WREA are summarized 

below. 

To assess the effects of visible foliar injury on recreation, the WREA reviewed the 

NSRE, as well as the 2006 National Survey of Fishing, Hunting, and Wildlife-Associated 

Recreation (FHWAR) and a 2006 analysis done for the Outdoor Industry Foundation (OIF). 

According to the NSRE, some of the most popular outdoor activities are walking, including day 

hiking and backpacking; camping; bird watching; wildlife watching; and nature viewing. 

Participant satisfaction with these activities can depend on the quality of the natural scenery, 

which can be adversely affected by O3-related visible foliar injury. According to the FHWAR 

and the OIF reports, the total expenditures across wildlife watching activities, trail-based 

activities, and camp-based activities are approximately $230 billion dollars annually. While the 

WREA could not quantify the magnitude of the impacts of O3 damage to the scenic beauty and 

outdoor recreation, the existing losses associated with current O3-related foliar injury are 

reflected in reduced outdoor recreation expenditures (U.S. EPA, 2014a, section 7.1). 

The WREA also assessed O3 impacts on cultural ecosystem services related to foliar 

injury at three national parks – Great Smoky Mountains National Park, Rocky Mountain 

National Park, and Sequoia/Kings National Parks - by considering information on visitation 

patterns, recreational activities and visitor expenditures. The analysis included percent cover of 

species sensitive to foliar injury and focused on the overlap between recreation areas within the 

park and elevated W126 concentrations. All three of these park units are in areas that are known 
5-68 

to have high ambient O3 concentrations, have vegetation maps, and have species that are 

considered O3-sensitive. Using GIS, the NPS vegetation maps were compared to the national O3

surface to illustrate where foliar injury may be occurring, particularly with respect to park 

amenities such as trails (U.S. EPA, 2014a, section 7.4). 

Great Smoky Mountains National Park is prized, in part, for its rich species diversity. 

The large mix of species includes 37 O3-sensitive species and many areas contain several 

sensitive species. With 3.8 million hikers using the trails every year and hikers’ WTP over $266 

million for that activity, even a small benefit of reducing O3 damage in the park could result in a 

significant economic value. Ozone concentrations in Great Smoky Mountains National Park 

have been among the highest in the eastern U.S. – at times twice as high as neighboring cities 

such as Atlanta (U.S. EPA, 2014a, p. 7-52). Unlike Great Smoky Mountains National Park, 

sensitive species cover in Rocky Mountain National Park is driven by a few O3-sensitive species 

(7 species) and most notably by Quaking Aspen. This is significant in that many of the visitors 

to Rocky Mountain National Park visit specifically to see this tree in its fall foliage. Given 1.5 

million hikers in Rocky Mountain National Park and their $70 million WTP for the hiking 

experience, even a small improvement in the scenic value could be economically significant 

(U.S. EPA, 2014a, section 7.4.2, p. 7-56). Sequoia/Kings National Parks is home to 12 identified 

sensitive species. Again, although the EPA is not able to quantify the impact of this scenic 

damage on hiker satisfaction for hikers in Sequoia/Kings National Parks and their $26 million 

WTP for the experience, even a small improvement in the scenic value could be economically 

significant ((U.S. EPA, 2014a, section 7.4.3, p. 7-63). 

 What are the uncertainties associated with this information and what is the level 

of confidence associated with those estimates? 

Uncertainties associated with these analyses are discussed in the WREA, sections 7.5 and 

8.5.3, and in WREA Table 7-24. As discussed in the WREA (section 8.5.3), evaluating soil 

moisture is more subjective than evaluating O3 exposure because of its high spatial and temporal 

variability within the O3 season, and there is considerable subjectivity in the categorization of 

relative drought. The WREA generally concludes that the spatial and temporal resolution for the 

soil moisture data is likely to underestimate the potential of foliar injury that could occur in some 

areas. In addition, there is lack of a clear threshold for drought below which visible foliar injury 

would not occur. In general, low soil moisture reduces the potential for foliar injury, but injury 

could still occur, and the degree of drought necessary to reduce potential injury is not clear. Due 

to the absence of biosite injury data in the Southwest region and limited biosite data in the West 

and West North Central regions, the benchmarks applied may not be applicable to these regions. 

The WREA applied the benchmarks from the national-scale analysis to a screening-level analysis 
5-69 

of 214 national parks and case studies of three national parks. Therefore, uncertainties in the 

foliar injury benchmarks are propagated into the national park analyses. 

There are also important uncertainties in the estimated O3 concentrations for the different 

air quality scenarios evaluated (U.S. EPA, 2014a, section 8.5), as discussed earlier in this 

section. These uncertainties only apply to the national park case studies because these are the 

only foliar injury analyses that rely on the air quality scenarios, but any uncertainties in the air 

quality analyses are propagated into the those analyses. Additional uncertainties are associated 

with the national park case studies. Specifically, there is uncertainty inherent in survey estimates 

of participation rates, visitor spending/economic impacts, and willingness-to-pay. These surveys 

potentially double-count impacts based on the allocation of expenditures across activities but 

also potentially exclude other activities with economic value. In general, the national level 

surveys apply standard approaches, which minimize potential bias. Other sources of uncertainty 

are associated with the mapping, including park boundaries, vegetation species cover, and park 

amenities, such as scenic overlooks and trails. In general, the WREA concludes that there is 

high confidence in the park mapping (U.S. EPA, 2014, Table 7-24). 

5.5 OTHER WELFARE EFFECTS 

In addition to the welfare effects discussed in the previous sections, there is evidence of 

other O3 effects, such as those related to climate impacts that we consider here. In this section, 

the WREA national-scale analyses of the effects of insect damage to forests related to elevated 

O3 exposures are considered in section 5.5.1, and a case study-scale characterization of the effect 

community composition changes on forest susceptibility and fire regulation in California is 

considered in section 5.5.2. As above, these sections, where possible, consider the WREA 

information regarding risk remaining under adjusted conditions just meeting the current standard 

and associated uncertainties (U.S. EPA 2014a, section 8.5). Chapters 5, 6, and 7 of the WREA 

also qualitatively assessed additional ecosystem services, including regulating services such as 

hydrologic cycle and pollination; provisioning services such as commercial non-timber forest 

products; and cultural services with aesthetic and non-use values. The information associated 

with these latter effects is insufficient to inform the target protection of the standard. The effects 

of O3 on climate are also considered in section 5.5.3 below, drawing primarily on the evidence 

presented in the ISA (U.S. EPA 2013, chapter 10). 

5.5.1 Forest Susceptibility to Insect Infestation 

Ozone in ambient air can contribute to increased susceptibility of some forests to 

infestation by some chewing insects, including the southern pine beetle and western bark beetle 

(U.S. EPA 2013, chapter 9; U.S. EPA 2014a, sections 5.3.3 and 5.4). These infestations can 
5-70 

cause economically significant damage to tree stands and the associated timber production. The 

WREA described the potential impacts of this effect on timber markets (U.S. EPA 2014a, section 

5.4). In the short-term, the immediate increase in timber supply that results from the additional 

harvesting of damaged timber depresses prices for timber and benefits consumers. In the longerterm, the decrease in timber available for harvest raises timber prices, potentially benefitting 

producers. The USFS reports timber producers have incurred losses of about $1.4 billion 

(2010$), and wood-using firms have gained about $966 million, due to beetle outbreaks between 

1977 and 2004. It is not possible to attribute a portion of these impacts resulting from the effect 

of O3 on trees’ susceptibility to insect attack; however, the losses are embedded in the estimates 

cited and any welfare gains from decreased O3 would positively impact the net economic impact. 

However, it is important to note that CASAC clarified that spatial association is not causation, 

even though expert opinion relates O3-exposure to bark beetle infestation (Frey, 2014, p. 12). 

To provide some quantitative estimates related to insect infestation-related risks, the 

WREA reported the estimates of 3-year average W126 index values in areas estimated to be at 

risk of greater than 25% timber loss (high loss) due to pine beetle infestation. This was done for 

all six WREA air quality scenarios. For example, for the recent conditions scenario, 

approximately 57 percent of the at-risk area has W126 estimates above 15 ppm-hrs, with the 

percentage dropping to approximately five percent in the current standard scenario (U.S. EPA 

2014a, section 5.4). 

5.5.2 Fire Regulation 

Evidence indicates that fire regime regulation may also be negatively affected by O3

exposure (U.S. EPA 2013, chapter 9; U.S. EPA 2014a, section 5.3.3). For example, Grulke et al. 

(2008) reported various lines of evidence indicating that O3 exposure may contribute to southern 

California forest susceptibility to wildfires by increasing leaf turnover rates and litter, increasing 

fuel loads on the forest floor. According to the National Interagency Fire Center, in the U.S. in 

2010 over 3 million acres burned in wildland fires and an additional 2 million acres were burned 

in prescribed fires. From 2004 to 2008, Southern California alone experienced, on average, over 

4,000 fires per year burning, on average, over 400,000 acres per fire. The California Department 

of Forestry and Fire Protection (CAL FIRE) estimated that losses to homes due to wildfire were 

over $250 million in 2007 (CAL FIRE, 2008). In 2008, CAL FIRE’s costs for fire suppression 

activities were nearly $300 million (CAL FIRE, 2008). 

The WREA developed maps that overlay the mixed conifer forest area of California with 

areas of moderate or high fire risk defined by CAL FIRE and with recent W126 concentrations 
5-71 

and surfaces adjusted to just meet existing and alternative standards. The highest fire risk and 

highest O3 concentrations overlap with each other, as well as with significant portions of mixed 

conifer forest. In the recent concentrations scenario, over 97 percent of mixed conifer forest area 

has average W126 index values over 7 ppm-hrs with a moderate to severe fire risk, and 74 

percent has average W126 index values over 15 ppm-hrs with a moderate to severe fire risk. The 

scenario for air quality adjusted to just meet the current standard, almost all of the mixed conifer 

forest area with a moderate to high fire risk shows a reduction in O3 to below a W126 index 

value of 7 ppm-hrs (average across three years of scenario). In the scenario for an average W126 

index value of 15 ppm-hrs, all but 0.18 percent of the area has average index values below 7 

ppm-hrs, and for the W126 scenarios of 11 and 7 ppm-hrs, all of the moderate to high fire threat 

area has estimated average W126 index values below 7 ppm-hrs (U.S. EPA 2014a, section 5.3.3, 

Figure 5-3). However, it is important to note that CASAC clarified that spatial association is not 

causation, but expert opinion relates O3-exposure to fire risk (Frey, 2014, p. 12). 

5.5.3 Ozone Effects on Climate 

Tropospheric O3 is a major greenhouse gas, third in importance after carbon dioxide 

(CO2) and methane (CH4). While the developed world has successfully reduced emissions of O3

precursors in recent decades, many developing countries have experienced large increases in 

precursor emissions and these trends are expected to continue, at least in the near term (U.S. 

EPA 2013, section 10.3.6.2). Projections of radiative forcing due to changing O3 over the 21st

century show wide variation, due in large part to the uncertainty of future emissions of source 

gases (U.S. EPA 2013, section 10.3.6.2). In the near-term (2000-2030), projections of O3

radiative forcing range from near zero to +0.3 W/m2, depending on the emissions scenario (U.S. 

EPA 2013, section 10.3.6.2; Stevenson et al., 2006). Reduction of tropospheric O3

concentrations could therefore provide an important means to slow climate change in addition to 

the added benefit of improving surface air quality (U.S. EPA, 2013, section 10.5). 

It is clear that increases in tropospheric O3 lead to warming. However the precursors of 

O3 also have competing effects on the greenhouse gas CH4, complicating emissions reduction 

strategies. A decrease in CO or VOC emissions would enhance OH concentrations, shortening 

the lifetime of CH4, while a decrease in NOX emissions could depress OH concentrations in 

certain regions and lengthen the CH4 lifetime (U.S. EPA, 2013, section 10.5). 

Abatement of CH4 emissions would likely provide the most straightforward means to 

address O3-related climate change since CH4 is itself an important precursor of background O3

(West et al., 2007; West et al., 2006; Fiore et al., 2002). A set of global abatement measures 

identified by West and Fiore (2005) could reduce CH4 emissions by 10% at a cost savings, 

decrease background O3 by about 1 ppb in the Northern Hemisphere summer, and lead to a 
5-72 

global net cooling of 0.12 W/m2

. West et al. (2007) explored further the benefits of CH4

abatement, finding that a 20% reduction in global CH4 emissions would lead to greater cooling 

per unit reduction in surface O3, compared to 20% reductions in VOCs or CO (U.S. EPA, 2013, 

section 10.5). 

Important uncertainties remain regarding the effect of tropospheric O3 on future climate 

change. To address these uncertainties, further research is needed to: (1) improve knowledge of 

the natural atmosphere; (2) interpret observed trends of O3 in the free troposphere and remote 

regions; (3) improve understanding of the CH4 budget, especially emissions from wetlands and 

agricultural sources, (4) understand the relationship between regional O3 radiative forcing and 

regional climate change; and (5) determine the optimal mix of emissions reductions that would 

act to limit future climate change (U.S. EPA, 2013, section 10.5). 

The IPCC has estimated the effect of the tropospheric O3 change since preindustrial times 

on climate to be about 25-40% of the anthropogenic CO2 effect and about 75% of the 

anthropogenic CH4 effect (IPCC, 2007). There are large uncertainties in the radiative forcing 

estimate attributed to tropospheric O3, making the effect of tropospheric O3 on climate more 

uncertain than the effect of the long-lived greenhouse gases (U.S. EPA, 2013, section 10.5). 

Radiative forcing does not take into account the climate feedbacks that could amplify or 

dampen the actual surface temperature response. Quantifying the change in surface temperature 

requires a complex climate simulation in which all important feedbacks and interactions are 

accounted for. As these processes are not well understood or easily modeled, the surface 

temperature response to a given radiative forcing is highly uncertain and can vary greatly among 

models and from region to region within the same model (U.S. EPA, 2013, section 10.5). 

As discussed in section 5.2 above, O3 exposure is associated with reduced forest tree 

growth, productivity, and carbon storage. Therefore, reducing O3 exposure would potentially 

increase carbon storage in O3-sensitive trees, which could also have climate effects. 

5.5.4 Additional Effects 

Recent information available since the last review considers the effects of O3 on chemical 

signaling in insect and wildlife interactions. Specifically, studies on O3 effects on pollination and 

seed dispersal, defenses against herbivory and predator-prey interactions all consider the ability 

of O3 to alter the chemical signature of VOCs emitted during these pheromone-mediated events. 

The effects of O3 on chemical signaling between plants, herbivores and pollinators as well as 

interactions between multiple trophic levels is an emerging area of study that may result in 

further elucidation of O3 effects at the species, community and ecosystem-level (U.S. EPA, 2013, 

p. 9-98). 
5-73 

5.6 CASAC ADVICE 

This section discusses CASAC advice regarding the adequacy of the existing secondary 

standard with respect to the 2008 review, the 2010 reconsideration of the 2008 review, and most 

recently its advice in this review, initiated in September 2008, in its letter to the Administrator on 

the second draft WREA and PA. To give an overview, following the 2008 decision to revise the 

secondary standard by setting it identical to the revised primary standard, CASAC conveyed 

additional advice to the Administrator regarding that decision. Shortly after that, several 

petitioners filed suit challenging the decision and in September 2009, the EPA announced its 

intention to reconsider the 2008 standards, issuing a notice of proposed rulemaking in January 

2010 (75 FR 2938). Soon after, the EPA solicited CASAC review of that proposed rule and in 

January 2011 solicited additional advice. This proposal was based on the scientific and technical 

record from the 2008 rulemaking, including public comments and CASAC advice and 

recommendations. As further described in section 1.2.2 above, the EPA in the fall of 2011 did 

not promulgate final rulemaking in that process but decided to coordinate further proceedings on 

the reconsideration rulemaking with this ongoing periodic review. 

More specifically, in April 2008, the members of the CASAC Ozone Review Panel sent a 

letter to EPA stating that “[i]n our most-recent letters to you on this subject - dated October 

2006 and March 2007 - ... the Committee recommended an alternative secondary standard of 

cumulative form that is substantially different from the primary Ozone NAAQS in averaging 

time, level and form — specifically, the W126 index within the range of 7 to 15 ppm-hours, 

accumulated over at least the 12 ‘daylight’ hours and the three maximum ozone months of the 

summer growing season” (Henderson, 2008). The letter continued: 

The CASAC now wishes to convey, by means of this letter, its additional, 

unsolicited advice with regard to the primary and secondary Ozone NAAQS. In 

doing so, the participating members of the CASAC Ozone Review Panel are 

unanimous in strongly urging you or your successor as EPA Administrator to 

ensure that these recommendations be considered during the next review cycle for 

the Ozone NAAQS that will begin next year ... The CASAC was also greatly 

disappointed that you failed to change the form of the secondary standard to 

make it different from the primary standard. As stated in the preamble to the 

Final Rule, even in the previous 1996 ozone review, “there was general 

agreement between the EPA staff, CASAC, and the Administrator, ... that a 

cumulative, seasonal form was more biologically relevant than the previous 1-

hour and new 8-hour average forms (61 FR 65716)” for the secondary 

standard......Unfortunately, this scientifically-sound approach of using a 

cumulative exposure index for welfare effects was not adopted... 
5-74 

In response to the EPA’s solicitation of their advice on the Agency’s proposed 

rulemaking as part of the reconsideration, CASAC conveyed their support for the proposed 

approach as follows (Samet, 2010). 

CASAC also supports EPA’s secondary ozone standard as proposed: a new 

cumulative, seasonal standard expressed as an annual index of the sum of 

weighted hourly concentrations (i.e., the W126 form), cumulated over 12 hours 

per day (8am to 8pm) during the consecutive 3-month period within the ozone 

season with the maximum index value, set as a level within the range of 7 to [1]5 

ppm-hours. This W126 metric can be supported as an appropriate option for 

relating ozone exposure to vegetation responses, such as visible foliar injury and 

reductions in plant growth. We found the Agency’s reasoning, as stated in the 

Federal Register notice of January 19, 2010, to be supported by the extensive 

scientific evidence considered in the last review cycle. In choosing the W126 form 

for the secondary standard, the Agency acknowledges the distinction between the 

effects of acute exposures to ozone on human health and the effects of chronic 

ozone exposures on welfare, namely that vegetation effects are more dependent on 

the cumulative exposure to, and uptake of, ozone over the course of the entire 

growing season (defined to be a minimum of at least three months). 

In its advice offered early in the current review, based on the updated scientific and 

technical record since the 2008 rulemaking, CASAC indicated that a conclusion that the current 

standard is inadequate to protect vegetation and ecosystems is “warranted” although it stated that 

the foundation needs to be broadened beyond analysis focused on Class I areas and trees to 

include “effects on sensitive crops, trees in regions outside of Class I areas, and additional 

ecosystem impacts” (Frey and Samet, 2012, p. 2). The Panel additionally endorsed the first draft 

PA discussions and conclusions on biologically relevant exposure metrics, stating that “the focus 

on the W126 form is appropriate” and that “there is a strong justification made for using a 

cumulative and weighted exposure standard for welfare effects (i.e., the W126)...” (Frey and 

Samet, 2012, p. 2). 

In its letter dated June 26, 2014, CASAC again concluded that “the current secondary 

standard is not adequate to protect against current and anticipated welfare effects of ozone on 

vegetation...” (Frey, 2014, p. iii) and that “the form of the standard should be changed from the 

current 8-hour form to the cumulative W126 index and... that the discussion provides an 

appropriate and sufficient rationale” (Frey, 2014, p. 12). CASAC then further states that 

“[t]hus, based on identification of known or anticipated ozone effects that are 

adverse to public welfare, taking into account the weight of evidence for causality of 

exposure to ozone and adverse welfare effects as given in Table 2-4 of the Integrated 

Science Assessment; results of the Second Draft WREA with regard to assessment of relative 
5-75 

biomass loss for tree species, foliar injury, and crop yield loss; and the breadth of adverse 

welfare effects for ecosystem services, foliar injury, and crop loss, the CASAC recommends 

that the secondary standard for ozone be revised as follows: (1) ozone should be the 

indicator; (2) the form and summation time of the standard should be the W126 index 

summed over the highest three-month interval during a year, based on accumulation over the 

08:00 a.m. – 08:00 p.m. daytime 12-hour period; and (3) the level of the standard should be 

between 7 ppm-hrs and 15 ppm-hrs. These recommendations are based on scientific 

evidence of adverse effect associated with the presence of ozone in ambient air. Note that 

these levels are based on an annual form of the standard.” (Frey, 2014, p. 15). 

With respect to the averaging time, CASAC additionally states that it “does not 

recommend the use of a three-year averaging period. We favor a single-year averaging 

period, which will provide more protection for annual crops and for the anticipated 

cumulative effects on perennial species. The scientific analyses considered in this review, 

and the evidence upon which they are based, are from single-year results. If a 3-year 

averaging period is established, then the upper limit will need to be reduced to protect 

against one-year ozone peaks” (Frey, 2014, p. 13). 

5.7 STAFF CONCLUSIONS ON ADEQUACY OF SECONDARY STANDARD 

This section presents staff conclusions for the Administrator to consider in deciding 

whether the existing secondary O3 standard is adequate and whether it should be retained or 

revised. Our conclusions are based on consideration of the assessment and integrative synthesis 

of information presented in the ISA, as well as our analyses of air quality distributions; analyses 

in the WREA; and the comments and advice of CASAC and public comment on earlier drafts of 

this document and on the ISA and WREA, as discussed above. Taking into consideration the 

responses to specific questions discussed above, we revisit the overarching policy question for 

this chapter: 

 Does the currently available scientific evidence and exposure/risk information, as 

reflected in the ISA and WREA, support or call into question the adequacy and/or 

appropriateness of the protection afforded by the current secondary O3 standard? 

As an initial matter, we note that the CAA does not require that a secondary standard be 

protective of all effects associated with a pollutant in the ambient air, but only those considered 

adverse to the public welfare (as described in section 1.3.2 above). In helping inform the 

Administrator’s judgments with respect to the adversity of the effects to public welfare, we have 

considered the scientific evidence and risk/exposure information in light of the paradigm used in 

the last review that takes into account the variation in public welfare significance of O3-related 
5-76 

vegetation effects when evaluating the potential adversity of the currently available evidence. As 

discussed in Section 5.1, this paradigm recognized that the significance to the public welfare of 

O3-induced effects on sensitive vegetation growing within the U.S. can vary depending on the 

nature of the effect, the intended use of the sensitive plants or ecosystems, and the types of 

environments in which the sensitive vegetation and ecosystems are located. Accordingly, any 

given O3-related effect on vegetation and ecosystems (e.g., biomass loss, crop yield loss, visible 

foliar injury) may be judged to have a different degree of impact on the public welfare 

depending, for example, on whether that effect occurs in a Class I area, a city park, or 

commercial cropland. In the 2010 proposed reconsideration, the Administrator proposed to place 

the highest priority and significance on vegetation and ecosystem effects to sensitive species that 

are known to or are likely to occur in federally protected areas such as national parks and other 

Class I areas, or on lands set aside by states, tribes and public interest groups to provide similar 

benefits to the public welfare (75 FR 3023/24), recognizing that effects occurring in such areas 

would likely have the highest potential for being classified as adverse to the public welfare, due 

to the expectation that these areas need to be maintained in a more pristine condition to ensure 

their intended use is met. 

In addition, there is also sufficient support to explicitly include consideration of impacts 

to ecosystem goods and services. Although ecosystem services were not explicitly considered in 

the Administrator’s decision in the last review, they were recognized as an important category of 

public welfare effects (73 FR 16492). The CASAC letter also provides support for this approach. 

The inclusion of ecosystem goods and services in this paradigm brings with it a number of 

additional considerations. Specifically, when considering the public welfare benefits from these 

goods and services, it is important to note that they can accrue across a range of dimensions, 

including spatial, temporal, and social, and these likely will vary depending on the type of effect 

being characterized. For example, ecosystems can cover a range of spatial scales, and the 

services they provide can accrue locally or be distributed more broadly such as when crops are 

sold and eaten locally and/or also sold in regional, national and world markets. Ecosystem 

services can likewise be realized over a range of temporal scales from immediate up to long term 

(e.g., the removal of air pollutants that have a short-term impact on human health but are also 

climate forcers with long atmospheric lifetimes, which the removal of may have immediate as 

well as long-term benefits). The size of the societal unit receiving benefits from ecosystem 

services can also vary dramatically. For example, a national park can provide direct recreational 

services to the thousands of visitors that come each year, but also provide an indirect value to the 

millions who may not visit but receive satisfaction from knowing it exists and is preserved for 

the future (U.S. EPA, 2014a, chapter 5, section 5.5.1). 
5-77 

We thus recognize the usefulness of evaluating the scientific evidence regarding these 

effects in the context of the most recent paradigm discussed above. This paradigm integrates the 

concepts of: 1) variability in public welfare significance given intended use and value of the 

affected entity, such as individual species; 2) relevance of associated ecosystem services to 

public welfare; and 3) variability in spatial, temporal, and social distribution of ecosystem 

services associated with known and anticipated welfare effects. In so doing, we recognize that 

there is no bright-line rule delineating the set of conditions or scales at which known or 

anticipated effects become adverse to public welfare. Thus, the evidence and exposure/risk 

information discussed in this chapter will be further evaluated in Chapter 6 in light of the 

concepts incorporated in this paradigm to help inform the Administrator’s judgments with 

respect to the potential adversity of the effects to the public welfare. 

With respect to the scientific evidence, the longstanding evidence base on the phytotoxic 

effects of O3 demonstrates that O3 -induced effects that occur at the subcellular and cellular 

levels, at sufficient magnitudes propagate up to larger spatial scales. The ISA summarizes the 

coherence across the full range of effects, from the least serious to the most serious, as follows 

(U.S. EPA, 2013, p. 1-8): 

The welfare effects of O3 can be observed across spatial scales, starting at the 

subcellular and cellular level, then the whole plant and finally, ecosystem-level 

processes. Ozone effects at small spatial scales, such as the leaf of an individual 

plant, can result in effects along a continuum of larger spatial scales. These 

effects include altered rates of leaf gas exchange, growth, and reproduction at the 

individual plant level, and can result in broad changes in ecosystems, such as 

productivity, carbon storage, water cycling, nutrient cycling, and community 

composition. 

Many of the recent studies evaluated in this review have focused on and further increased 

our understanding of the molecular, biochemical and physiological mechanisms that explain how 

plants are affected by O3, in the absence of other stressors, particularly in the area of genomics 

(U.S. EPA, 2013, Chapter 9, section 9.3). These recent studies, in combination with the 

extensive and long-standing evidence, have further strengthened the coherence and consistency 

of the entire body of research, so that our confidence in the supporting science is stronger than in 

the previous review. 

Based on its assessment of the strength of the science, the ISA determined that the 

relationship that exists between exposure to O3 in ambient air and visible foliar injury effects on 

vegetation, reduced vegetation growth, reduced productivity in terrestrial ecosystems, reduced 

yield and quality of agricultural crops and alteration of below-ground biogeochemical cycles 

(U.S. EPA 2013, Table 1-2) is causal. Additionally, the ISA determined that a likely to be causal 
5-78 

relationship exists between exposures to O3 in ambient air and reduced carbon sequestration in 

terrestrial ecosystems, alteration of terrestrial ecosystem water cycling and alteration of 

terrestrial community composition (U.S. EPA, 2013, Table 1-2). 

Recent studies also continue to provide strong and consistent evidence that adverse 

vegetation effects are attributable to cumulative seasonal O3 exposures. On the basis of the 

entire body of evidence in this regard, the ISA concludes that “quantifying exposure with indices 

that cumulate hourly O3 concentrations and preferentially weight the higher concentrations 

improves the explanatory power of exposure/response models for growth and yield, over using 

indices based on mean and peak exposure values” (U.S. EPA, 2013, p. 2-44). Thus, as in other 

recent reviews, the evidence continues to provide a strong basis for concluding that it is 

appropriate to judge impacts of O3 on vegetation, related effects and services, and the level of 

public welfare protection achieved, using a cumulative, seasonal exposure metric, such as the 

W126-based metric. In addition, CASAC has consistently since the 1997 review expressed 

support for the use of such a metric as the most appropriate form for the secondary NAAQS. In 

its most recent letter on the second draft PA, CASAC states that it “concurs with the justification 

in this section that the form of the standard should be changed from the current 8-hr form to the 

cumulative W126 index and finds that the discussion provides an appropriate and sufficient 

rationale” (Frey, 2014, p. 12). Thus, based on the consistent and well-established evidence 

described above, we conclude that the most appropriate and biologically relevant way to relate 

O3 exposure to plant growth, and to determine what would be adequate protection for public 

welfare effects attributable to the presence of O3 in the ambient air, is to characterize exposures 

in terms of a cumulative seasonal form, and in particular the W126 metric. 

 Accordingly, in considering the current evidence and exposure/risk information with 

regard to the adequacy of public welfare protection it affords, we have considered both the 

evidence of vegetation and welfare impacts in areas of the U.S. likely to have met the current 

standard, as well as air quality information regarding W126 index values in such areas. In 

evaluating the adequacy of the current secondary standard, we first considered O3 effects on tree 

growth, productivity and carbon storage and associated ecosystems and services. Recent studies 

confirm and extend the evidence of O3-related effects on tree growth, productivity and carbon 

storage. Analysis of existing data conducted by the EPA staff and discussed in the ISA has 

substantially reduced the uncertainty associated with using OTC E-R functions to predict tree 

growth effects in the field, as described in section 5.2.1 above (U.S. EPA, 2013, section 9.6.3.2). 

The median of the composite E-R functions (green line), (U.S. EPA, 2014a, Figure 6-5, section 

6.2.1.2) shows RBL for tree seedlings. We note CASAC’s advice that a 6% median RBL is 

unacceptably high, and that the 2% median RBL is an important benchmark to consider. The 
5-79 

median RBL is at or below 2% at the lowest W126 level assessed, 7 ppm-hrs. As the W126 

level is incrementally increased, median RBL also increases incrementally, so that at W126 

index values of 9, 11, 13, 15, 17, 19 and 21, the median RBL increases to 2.4%, 3.1%, 3.8%, 

4.5%, 5.3%, 6.0% and 6.8%, respectively. Based on air quality analyses of 2009-2011 (Appendix 

2B), there are approximately 342, 199, 92, 43, 24, 9, 3 and 0 monitors with 3-year average W126 

index values above 7, 9, 11, 13, 15, 17, 19 and 21 ppm-hrs when meeting the current standard. 

We note that these counts of monitors are based on those meeting the current standard and that 

there are many monitors for the 2009-2011 period that do not meet the current standard and also 

are above the W126 values of 7-21 ppm-hrs. 

 We also consider it informative to examine the individual species responses and RBL 

over the same W126 range. We first note, based on Figure 5-1 (B) above that over the range of 7 

to 17 ppm-hrs, 5 species maintain RBLs of less than 2%. These more tolerant species include 

Douglas fir, loblolly pine, Virginia pine, sugar maple and red maple. Two of these species (red 

maple and sugar maple) are estimated to have RBL levels above 2% at a W126 of 21. Black 

cherry, the most sensitive of the remaining six species, has RBL ranging from 35.57% at W126 

of 17 down to 16.67% at the W126 index value of 7 ppm-hrs. 

In addition, we also consider the growth effects associated with exposure concentrations 

at or below that of the current standard in Class I areas. Specifically, we found that there were 22 

Class I areas that had monitor sites that have design values that meet the current standard, 

ranging from 67 to 75 ppb, and have 3-year average W126 index values that are above 15 ppmhrs between the years of 1998 and 2012 (Table 5-2). Across these 22 Class I areas, the highest 

single-year W126 index values for these three-year periods ranged from 17.4 to 29.0 ppm-hrs. In 

20 of the areas, distributed across eight states (AZ, CA, CO, KY, NM, SD, UT, WY) and four 

regions (west, southwest, west/north central and central), this range was 19.1 to 29.0 ppm-hrs, 

exposure values for which the corresponding median species RBL estimates equal or exceed 6%, 

which CASAC termed “unacceptably high”. In addition, given that other environmental factors 

can influence the extent to which O3 may have the impact predicted by the E-R functions in any 

given year, we also note that the highest three year periods, that include these highest annual 

values for the 21 areas, are at or above 19 ppm-hrs, ranging up to 22.5 ppm-hrs (for which the 

median species RBL estimate is above 7%). Additionally, the highest three-year average W126 

index value for each of the 22 areas (during periods meeting the current standard) was at or 

above 19 (ranging up to 22.5 ppm-hrs) in 11 areas, distributed among five states in the west and 

southwest regions (U.S. EPA, 2014c, Table 5-2, Appendix 5B). 

In addition, quaking aspen and ponderosa pine are two tree species that are found in most 

of these 22 parks and have a sensitivity to O3 exposure that places them near the middle of the 
5-80 

group for which E-R functions have been established. In the areas where ponderosa pine is 

present, the highest single year W127 index values ranged from 18.7 to 29.0 and the highest 3-

year average W126 values in which these single year values are represented ranged from 15 to 

22.5, with these three-year values above 19 ppm-hrs in eight areas across five states. The 

ponderosa pine RBL estimates for 29 and 22.5 ppm-hrs are approximately 12% and 9%, 

respectively. In the areas where quaking aspen is present, the highest single year W127 index 

values ranged from 19.2 to 26.7 ppm-hrs and the highest 3-year average W126 values in which 

these single year values are represented ranged from 15.0 to 22.2, with values above 19 ppm-hrs 

in eight areas across five states. The quaking aspen RBL estimates for 26.7 and 22.2 ppm-hrs are 

approximately 16% and 13%, respectively. Based on this, we predict growth effects associated 

with exposure concentrations at or below that of the current standard for most of these Class I 

areas. On the basis of such information, Table 5-2 provides evidence of the potential for 

significant growth loss in locations where ambient conditions meet the current standard. Based 

on this evidence, we note the occurrence in Class I areas, during periods where the current 

standard is met, of cumulative seasonal O3 exposures of a magnitude that might reasonably be 

concluded to be important to public welfare. 

Recent studies have provided additional evidence on tree biomass or growth effects 

associated with multiple year exposures in the field, including the potential for cumulative and 

carry-over effects. For example, a number of studies were conducted at a planted forest at the 

Aspen FACE site in Wisconsin where some researchers observed that the effects of O3 on birch 

seeds (reduced weight, germination, and starch levels) could lead to a negative impact on species 

regeneration in subsequent years, and that the effect of reduced aspen bud size may have been 

related to the observed delay in spring leaf development. These effects suggest that elevated O3

exposures have the potential to alter carbon metabolism of overwintering buds which may have 

subsequent effects in the following year. Other studies found that, in addition to affecting tree 

heights, diameters, and main stem volumes in the aspen community, elevated O3 over a 7-year 

study period was reported to increase the rate of conversion from a mixed aspen-birch 

community to a community dominated by the more tolerant birch, leading the authors to 

conclude that elevated O3 may alter intra- and inter-species competition within a forest stand 

(U.S. EPA, 2013, section 9.4.3). 

While it is not possible at this time to identify the extent or magnitude of such effects in 

the field under exposures that may be associated with the current standard, their occurrence, on 

federal lands with special protections might reasonably be concluded to be an important public 

welfare consideration. We note here that the CASAC “concurs that biomass loss in trees is a 

relevant surrogate for damage to tree growth that affects ecosystem services such as habitat 

provision for wildlife, carbon storage, provision of food and fiber, and pollution removal. 
5-81 

Biomass loss may also have indirect process-related effects such as on nutrient and hydrologic 

cycles. Therefore, biomass loss is a scientifically valid surrogate of a variety of adverse effects to 

public welfare” (Frey, 2014, pp. 9-10). 

In regard to the WREA analyses for risks for associated ecosystem services, we note that 

the WREA presents estimated changes in consumer and producer/farmer surplus associated with 

the change in forestry and agricultural yields. Changes in biomass affect individual tree species 

differently, and the overall effect on forest ecosystem productivity depends on the composition 

of forest stands and the relative sensitivity of trees within those stands. Economic welfare 

impacts resulting from just meeting the existing and alternative standards were largely similar 

between the forestry and agricultural sectors -- consumer surplus, or consumer gains, generally 

increased in both sectors because higher productivity under lower O3 concentrations increased 

total yields and reduced market prices. Comparisons are not straightforward to interpret due to 

market dynamics. The national-scale analysis of carbon dioxide (CO2) sequestration estimates 

more storage under the current standard compared to recent conditions, with somewhat smaller 

additional increases for the three W126 scenarios in comparison to the current standard scenario 

(U.S. EPA 2014a, Appendix 6B, Table B-10). 

We additionally consider the WREA estimates of tree growth and ecosystem services 

provided by urban trees over a 25-year period for five urban areas based on case-study scale 

analyses that quantified the effects of biomass loss on carbon sequestration and pollution 

removal (U.S. EPA 2014a, sections 6.6.2 and 6.7).30 The urban areas included in this analysis 

represent diverse geography in the Northeast, Southeast, and Central regions, although they do 

not include an urban area in the western U.S. Estimates of the effects of O3-related biomass loss 

on carbon sequestration, for example, indicate the potential for an increase of somewhat more 

than a million metric tons of CO2 equivalents for average W126 index values associated with 

meeting the current standard scenario as compared to recent conditions. Somewhat smaller 

additional increases are estimated for the three W126 scenarios in comparison to the current 

standard scenario (U.S. EPA 2014a, section 6.6.2 and Appendix 6D). 

In considering the significance of these WREA analyses of risks for the associated 

ecosystem services for timber production, air pollution removal, and carbon sequestration, we 

note the large uncertainties associated with these analyses (see U.S. EPA 2014a, Table 6-27), and 

the potential to underestimate the response at the national scale. Thus, while we note that it is 

appropriate to consider predicted and anticipated impacts to these services in determining the 

adequacy of the protection afforded by the current standard, we also note that we place limited 

 

30 The WREA used the i-Tree model for the urban case studies. i-Tree is a peer-reviewed suite of software 

tools provided by USFS. 
5-82 

weight on the absolute magnitude of the risk results for these ecosystem service endpoints in 

light of these significant associated uncertainties. 

In reaching conclusions regarding support for the adequacy of the current secondary 

standard provided by the currently available information on O3-induced effects on trees and 

associated ecosystem services we note that: 1) there is robust evidence supporting the causal 

relationship between cumulative O3 exposures and effects on tree growth, productivity, and 

carbon storage (U.S. EPA, 2013) and causal and likely to be causal relationships for several 

associated ecosystem services; 2) the tree seedling E-R functions evidence, which has been 

strengthened, demonstrates variability in sensitivity to O3 across species; 3) estimated median 

RBLs are at or above 6%, a key CASAC benchmark, in several areas when air quality was at or 

below that of the current standard; 4) growth effects associated with exposure concentrations are 

predicted to occur in several Class I areas based on air quality from 1998-2012 that was at or 

below that of the current standard; 5) impacts from single year exposures can carry over to the 

subsequent year and/or cumulate over multiple years with repeated annual exposures; 6) 

evidence from both recent controlled chamber mechanism studies and field based exposure 

studies support earlier findings from OTC studies; and 7) WREA analyses show that O3-induced 

biomass loss can impact ecosystem services provided by forests, including timber production, 

carbon storage, and air pollution removal, even when air quality is adjusted to just meet the 

current standard. Given the above, and noting CASAC views described above, staff concludes 

that the current evidence/risk information calls into question the adequacy of the public welfare 

protection afforded by the current standard from the known and anticipated adverse effects 

associated with O3-induced impacts on tree growth, productivity and carbon storage, including 

the associated ecosystem services assessed in this review, and therefore it is appropriate to 

consider revision to provide increased protection. 

With respect to crops, the detrimental effect of O3 on crop production has been 

recognized since the 1960s, and recent O3 concentrations in many areas across the U.S. are high 

enough that they might be expected to cause yield loss in a variety of agricultural crops 

including, but not limited to, soybeans, wheat, potatoes, watermelons, beans, turnips, onions, 

lettuces, and tomatoes (U.S. EPA, 2013, section 9.4.4). In general, the vast majority of the new 

scientific information confirms prior conclusions that exposure to O3 can decrease growth and 

yield of crops. Recent research has highlighted the effects of O3 on crop quality. Increasing O3

concentration decreases nutritive quality of grasses, and decreases macro- and micro-nutrient 

concentrations in fruits and vegetable crops (U.S. EPA 2013, section 9.4.4). Recent studies 

continue to find yield loss levels in crop species studied previously under NCLAN that reflect 

the earlier findings. There has been little published evidence that crops are becoming more 

tolerant of O3 (U.S. EPA, 2006a; U.S. EPA 2013). This is especially evident in the research on 
5-83 

soybean. The 2013 ISA reported comparisons between yield predictions based on data from 

cultivars used in NCLAN studies, and yield data for modern cultivars from SoyFACE (U.S. 

EPA, 2013, section 9.6.3). They confirm that the average response of soybean yield to O3

exposure has not changed in current cultivars. In addition, satellite and ground-based O3 

measurements have been used to assess yield loss caused by O3 over the continuous tri-state area 

of Illinois, Iowa, and Wisconsin. The results showed that O3 concentrations reduced soybean 

yield, which correlates well with the previous results from FACE- and OTC-type experiments 

(U.S. EPA, 2013, section 9.4.4.1). Thus, the recently available evidence, as assessed in the ISA, 

continues to support the conclusions of the 1996 and 2006 CDs that ambient O3 concentrations 

can reduce the yield of major commodity crops in the U.S. 

The currently available evidence, as assessed in the ISA, continues to support the use of 

E-R functions for crops based on OTC experiments. Further, important uncertainties have been 

reduced regarding the E-R functions for crop yield loss, especially for soybean. In general, the 

ISA reports consistent results across exposure techniques and across crop varieties (U.S. EPA 

2013, section 9.6.3.2). Soybean, which is the second-most planted field crop in the U.S.,31 would 

be predicted to have no more than 5% RYL at a W126 index value of 12 ppm-hrs, based on the 

E-R function. Staff analyses of recent monitoring data (2009-2011) indicate that O3

concentrations in multiple agricultural areas in the U.S. that meet the current standard correspond 

to W126 index levels above 12 ppm-hrs. With regard to crops, CASAC states that it “concurs 

that another important surrogate for damage that is adverse to public welfare is crop loss. Crops 

provide food and fiber services to humans. Evaluation of market-based welfare effects of ozone 

exposure in forestry and agricultural sectors is an appropriate approach to take into account 

damage that is adverse to public welfare” (Frey, 2014, p. 10). However, as we describe in 

section 5.3 above, determining at what point O3-induced crop yield loss becomes adverse to the 

public welfare is still unclear, given that it is heavily managed with additional inputs that have 

their own associated markets and that benefits can be unevenly distributed between producers 

and consumers. We further note that a standard set to provide requisite protection for trees could 

also potentially achieve appropriate protection for commodity crops. 

In reaching conclusions regarding support for the adequacy of the current secondary 

standard provided by the currently available information on O3-induced effects on crops, we note 

that 1) there is clear and robust evidence supporting the causal relationship between cumulative 

O3 exposures and effects on crop yields and quality (U.S. EPA, 2013); 2) the crop E-R functions 

evidence, which has been strengthened, demonstrates variability in sensitivity to O3 across 

species; 3) evidence from both recent controlled chamber mechanism studies and field based 

 

31 http://www.ers.usda.gov/topics/crops/soybeans-oil-crops/background.aspx
5-84 

exposure studies support earlier findings from OTC studies; 4) evidence continues to show that 

crops, and in particular soybean, has not become more tolerant of O3 (U.S. EPA, 2013, section 

9.6.3, 9.4.4.1); 5) WREA analyses show that O3-induced crop yield loss can impact producer and 

consumer surpluses and the interaction between agriculture and timber production. 

Given the above, and noting CASAC views described above as well as the difficulty in 

assessing adversity to public welfare of these effects, staff concludes that the current 

evidence/risk information calls into question the adequacy of the public welfare protection 

afforded by the current standard from the known and anticipated adverse effects associated with 

O3-induced impacts on crop yields and associated services assessed in this review, and therefore, 

it is appropriate to consider revision to provide increased protection. 

With respect to foliar injury, visible foliar injury surveys are used by the federal land 

managers to assess potential O3 impacts in Class I areas (USFS, NPS, FWS, 2010). Given this 

focus on visible foliar injury, O3-induced impacts have the potential to impact the public welfare 

in scenic and/or recreational areas on an annual basis. Visible foliar injury is associated with 

important cultural and recreational ecosystem services to the public, such as scenic viewing, 

wildlife-watching, hiking, and camping, that are of significance to the public welfare and 

enjoyed by millions of Americans every year, generating millions of dollars in economic value 

(U.S. EPA 2014a, section 7.1). In addition, several tribes have indicated that many of the known 

confirmed O3-sensitive species (including bioindicator species) are culturally significant (see 

Appendix 5-A). We further note that CASAC “concurs that visible foliar injury can impact 

public welfare by damaging or impairing the intended use or service of a resource. Visible foliar 

injury that is adverse to public welfare can include: visible damage to ornamental or leafy crops 

that affects their economic value, yield, or usability; visible damage to plants with special 

cultural significance; and visible damage to species occurring in natural settings valued for 

scenic beauty or recreational appeal” (Frey, 2014, p. 10). 

New research on visible foliar injury includes: 1) controlled exposure studies; 2) multiyear field surveys; and 3) USFS FHM/FIA biomonitoring program surveys and assessments. In 

addition to supporting the ISA causal determination, these studies also address some 

uncertainties identified in the previous review (i.e., influence of soil moisture on visible foliar 

injury development) and further show that visible foliar injury can occur under conditions where 

8-hour average O3 concentrations are or would be expected to be below the level of the current 

standard (e.g., Kline et al., 2008, as discussed in section 5.4.1 above). Incidence of visible foliar 

injury symptoms on O3-sensitive vegetation has also been documented in the field in federally 

protected areas that have met the current standard. Importantly, these O3-induced vegetation 

effects have been identified by the federal land managers as a diagnostic tool for informing 

conclusions regarding potential ozone impacts on potentially sensitive AQRVs and were found 
5-85 

in Class I areas that have particular public welfare significance in light of direction from 

Congress that these areas as merit a high level of protection (75 FR 3023/3024). 

The studies mentioned above also provide additional information regarding the role of 

soil moisture in influencing visible foliar injury response (U.S. EPA 2013, section 9.4.2). These 

studies confirm that adequate soil moisture creates an environment conducive to greater visible 

foliar injury in the presence of O3 than drier conditions. As stated in the ISA, “[a] major 

modifying factor for O3-induced visible foliar injury is the amount of soil moisture available to a 

plant during the year that the visible foliar injury is being assessed ... because lack of soil 

moisture generally decreases stomatal conductance of plants and, therefore, limits the amount of 

O3 entering the leaf that can cause injury” (U.S. EPA, 2013, p. 9-39). As a result, “many studies 

have shown that dry periods in local areas tend to decrease the incidence and severity of O3-

induced visible foliar injury; therefore, the incidence of visible foliar injury is not always higher 

in years and areas with higher O3, especially with co-occurring drought (Smith, 2012; Smith et 

al., 2003)” (U.S. EPA, 2013, p. 9-39). This “...partial ‘protection’ against the effects of O3

afforded by drought has been observed in field experiments (Low et al., 2006) and modeled in 

computer simulations (Broadmeadow and Jackson, 2000)” (U.S. EPA, 2013, p. 9-87). In 

considering the extent of any protective role of drought conditions, however, the ISA also notes 

that other studies have shown that “drought may exacerbate the effects of O3 on plants 

(Pollastrini et al., 2010; Grulke et al., 2003)” and that “[t]here is also some evidence that O3 can 

predispose plants to drought stress (Maier-Maercker, 1998)”. Accordingly, the ISA concludes 

that “the nature of the response is largely species-specific and will depend to some extent upon 

the sequence in which the stressors occur” (U.S. EPA, 2013, p. 9-87). However, such 

uncertainties associated with describing the potential for foliar injury and its severity or extent of 

occurrence for any given air quality scenario due to confounding by soil moisture levels make it 

difficult to identify an appropriate degree of protection (as well as ambient O3 exposure 

conditions that might be expected to provide that protection). 

We note the WREA analyses of the nationwide dataset (2006-2010) for USFS FHM/FIA 

biosites described in section 5.4.2 above, including the observation that the proportion of biosites 

with injury varies with soil moisture conditions and O3 W126 index values (U.S. EPA 2014a, 

Chapter 7, Figure 7-10). The evidence of O3-attributable visible foliar injury incidence occurring 

in USFS FHM/FIA biosites shows that the proportion of biosites showing foliar injury incidence 

increases steeply with W126 index values up to approximately 10 ppm-hrs. At W126 index 

levels greater than approximately 10 ppm-hrs, the proportion of sites showing foliar injury 

incidence is relatively constant. The air quality assessment discussed above identified Class I 

areas with recent air quality that met the current standard but were above a W126 index value of 
5-86 

15 ppm-hrs (Table 5-2). There were 22 Class I areas in this table and most of these areas had 3-

year average W126 index values above 15 ppm-hrs for multiple 3-year periods. Given evidence 

of the potential occurrence of visible foliar injury at W126 index values of this magnitude, we 

note the ecosystem services that are at risk of impairment because of O3-induced damage directly 

due to foliar injury, though data is not available to explicitly quantify these negative effects. 

Therefore, staff concludes that air quality levels that are at or below the level of the current 

standard may allow levels of visible foliar injury incidence to occur in areas of special 

significance to the public welfare. 

In reaching conclusions regarding support for the adequacy of the current secondary 

standard provided by the currently available information on O3-induced visible foliar injury we 

note that: 1) many species of native plants, including trees, have been observed to have visible 

foliar injury symptoms in both OTC and field settings, some of which have also been identified 

as bioindicators of O3 exposure by the USFS; 2) visible foliar injury has been identified by the 

federal land managers as a diagnostic tool for informing conclusions regarding potential O3

impacts on potentially sensitive AQRVs (USFS, NPS, FWS, 2010); 3) visible foliar injury 

incidence can occur for some species at very low cumulative exposures, but due to confounding 

by soil moisture and other factors, it difficult to predictively relate a given O3 exposure to plant 

response; and 4) WREA analyses show that based on USFS biosite data, the proportion of 

biosites showing foliar injury incidence drops when W126 index values drop below 

approximately 10 ppm-hrs. However, we note that, with respect to visible foliar injury, we are 

unaware of any guidance for federal land managers regarding at what spatial scale or what 

degree of severity visible foliar injury might be sufficient to trigger protective action based on 

this potential impact on AQRVs. Further, there does not appear to be any consensus in the 

literature in this regard, and CASAC, while identifying target percent biomass loss and yield loss 

benchmarks for tree seedlings and commodity crops, respectively, did not provide a similar 

recommendation for this endpoint. Likewise, as in previous reviews, the ISA notes the difficulty 

in relating visible foliar injury symptoms to other vegetation effects such as individual plant 

growth, stand growth, or ecosystem characteristics (U.S. EPA, 2013, section 9.4.2, p. 9-39) and 

further noted that the full body of evidence indicates that there is wide variability in this 

endpoint, such that although evidence shows visible foliar injury can occur under very low 

cumulative O3 concentrations, “...the degree and extent of visible foliar injury development 

varies from year to year and site to site..., even among co-members of a population exposed to 

similar O3 levels, due to the influence of co-occurring environmental and genetic factors” (U.S. 

EPA 2013, section 9.4.2, p. 9-38). 

 Given the above, and taking note of CASAC views, we recognize foliar injury as an 

important O3 effect which, depending somewhat on severity and spatial extent, may reasonably 
5-87 

be concluded to be of public welfare significance, especially when occurring in nationally 

protected areas. However, we also note the uncertainties associated with describing the potential 

for foliar injury and its severity or extent of occurrence for any given air quality scenario due to 

confounding by soil moisture levels and the difficulty in determining what degree of visible 

foliar injury incidence is likely to occur under different air quality conditions, and in particular 

on lands with special public welfare significance. We therefore conclude that the current 

standard may not adequately protect the public welfare from the known and anticipated adverse 

effects associated with O3-induced impacts on visible foliar injury and associated services 

assessed in this review. Therefore, it may be appropriate to consider revising the standard to 

provide increased public welfare protection, though it is uncertain to what degree these O3-

induced impacts on visible foliar injury would be appropriately judged as important and adverse 

from a public welfare perspective. 

The information for other welfare effects, including those with causal or likely causal 

relationships with O3 (e.g., alteration of ecosystem water cycling, changes in climate), is limited 

with regard to our ability to consider potential impacts of air quality conditions associated with 

the current standard, although the WREA provides some perspective on this issue with regard to 

susceptibility to insect attack and fire regime change. We note, however, the importance of these 

effects categories to the public welfare. 

As noted in section 1.3.2 above, our general approach to informing the Administrator’s 

judgments recognizes that the available welfare effects evidence demonstrates a range of O3

sensitivity across studied plant species and documents an array of O3-induced effects that extend 

from lower to higher levels of biological organization. These effects range from those affecting 

cell processes and individual plant leaves to effects on the physiology of whole plants, species 

effects and effects on plant communities to effects on related ecosystem processes and services. 

Given this evidence, it is not possible to generalize across all studied species regarding which 

cumulative exposures are of greatest concern, as this can vary by situation due to differences in 

exposed species sensitivity, the importance of the observed or predicted O3-induced effect, the 

role that the species plays in the ecosystem, the intended use of the affected species and its 

associated ecosystem and services, the presence of other co-occurring predisposing or mitigating 

factors, and associated uncertainties and limitations. At the same time, the evidence also 

demonstrates that though effects of concern can occur at very low exposures in sensitive species, 

at higher cumulative exposures those effects would likely occur at a greater magnitude and/or 

higher levels of biological organization and additional species would likely be impacted. It is 

important to note, however, due to the variability in the importance of the associated ecosystem 

services provided by different species at different exposures and in different locations, as well as 

differences in associated uncertainties and limitations, that, in addition to the magnitude of the 
5-88 

ambient concentrations, the species present and their public welfare significance are essential 

considerations in drawing conclusions regarding the significance of public welfare impact. 

Therefore, in developing conclusions in this final PA, we note the complexity of 

judgments to be made by the Administrator regarding the adversity of known and anticipated 

effects to the public welfare and are mindful that the Administrator’s ultimate judgments on the 

secondary standard will most appropriately reflect an interpretation of the available scientific 

evidence and exposure/risk information that neither overstates nor understates the strengths and 

limitations of that evidence and information. 

Given all of the above, we reach the conclusion that the available evidence and exposure 

and risk information call into question the adequacy of public welfare protection provided by the 

current standard, and provides support for consideration of revisions to the current secondary 

standard to provide increased public welfare protection. More specifically, staff concludes that it 

is appropriate for the Administrator to consider revision of the current secondary O3 standard to 

increase protection against O3-attributable tree biomass loss, crop yield loss, and visible foliar 

injury, and their associated services, and particularly for those effects associated with 

cumulative, seasonal exposures that occur in Class I and similarly protected natural areas. 

In reaching conclusions on options for the Administrator’s consideration, we note that the 

final decision to retain or revise the current secondary O3 standard is a public welfare policy 

judgment to be made by the Administrator, based on her judgment as to what degree of 

protection would be requisite (i.e., neither more nor less stringent than necessary) to protect the 

public welfare from any known or anticipated adverse effects. This final decision will draw upon 

the available scientific evidence for O3-attributable welfare effects, and on quantitative analyses 

of vegetation and ecosystem exposures and associated risks to vegetation, ecosystems and their 

associated services, and judgments about the appropriate weight to place on the range of 

uncertainties inherent in the evidence and analyses. In making this decision, the Administrator 

will also need to weigh the importance of these effects and their associated ecosystem services in 

the overall context of public welfare protection. 

Based on the considerations described in the sections above and summarized below, we 

therefore conclude that the currently available evidence and exposure/risk information call into 

question the adequacy of the public welfare protection provided by the current standard and 

provides support for considering potential alternative standards to achieve increased public 

welfare protection, especially for sensitive vegetation and ecosystems in federally protected 

Class I and similar areas. In this conclusion, we give particular weight to the evidence indicating 

the occurrence in Class I areas that meet the current standard of cumulative seasonal O3

exposures associated with estimates of tree growth impacts of a magnitude that are reasonably 

considered important to public welfare. 
5-89 

5.8 REFERENCES 

Abt Associates, Inc. (1995). Ozone NAAQS benefits analysis: California crops. Report to U.S. EPA, July. 

Andersen, CP; Wilson, R; Plocher, M; Hogsett, WE. (1997). Carry-over effects of ozone on root growth and 

carbohydrate concentrations of ponderosa pine seedlings. Tree Physiol 17: 805-811. 

Betzelberger, AM; Gillespie, KM; McGrath, JM; Koester, RP; Nelson, RL; Ainsworth, EA. (2010). Effects of 

chronic elevated ozone concentration on antioxidant capacity, photosynthesis and seed yield of 10 soybean 

cultivars. Plant Cell Environ 33: 1569-1581. http://dx.doi.org/10.1111/j.1365-3040.2010.02165.x 

Bortier, K.; DeTemmerman, L.; Ceulemans, R. (2000b) Effects of ozone exposure in open-top chambers on poplar 

(Populus nigra) and beech (Fagus sylvatica): a comparison. Environ. Pollut. 109: 509-516.

Broadmeadow, MSJ; Jackson, SB. (2000). Growth responses of Quercus petraea, Fraxinus excelsior and 

Pinussylvestris to elevated carbon dioxide, ozone and water supply. New Phytol 146: 437-

451.http://dx.doi.org/10.1046/j.1469-8137.2000.00665.x 

Burns, Russell M., and Barbara H. Honkala, tech. coords. (1990). Silvics of North America: 1. Conifers; 2. 

Hardwoods. Agriculture Handbook 654. U.S. Department of Agriculture, Forest Service, Washington, DC. 

vol.2, 877 p. 

Burns, Russell M., and Barbara H. Honkala, tech. coords. (1990). Silvics of North America: 1.Conifers; 2. 

Hardwoods. Agriculture Handbook 654. U.S. Department of Agriculture, Forest Service, Washington, DC. 

vol.2, 877 p. 

CAL FIRE (California Department of Forestry and Fire Protection). (2008). CAL FIRE 2007 Wildland Fire 

Summary. Available at www.fire.ca.gov/communications/downloads/fact_sheets/2007Summary.pdf. 

Campbell, SJ; Wanek, R; Coulston, JW. (2007). Ozone injury in west coast forests: 6 years of monitoring - 

Introduction. Portland, OR: U.S. Department of Agriculture. 

Darbah, JNT; Kubiske, ME; Neilson, N; Oksanen, E; Vaapavuori, E; Karnosky, DF. (2007). Impacts of elevated 

atmospheric CO2 and O3 on paper birch (Betula papyrifera): Reproductive fitness. Scientific World Journal 

7: 240-246. http://dx.doi.org/10.1100/tsw.2007.42 

Darbah, JNT; Kubiske, ME; Nelson, N; Oksanen, E; Vapaavuori, E; Kamosky, DF. (2008). Effects of decadal 

exposure to interacting elevated CO2 and/or O-3 on paper birch (Betula papyrifera) reproduction. Environ 

Pollut 155: 446-452. http://dx.doi.org/10.1016/j.envpol.2008.01.033Davis, DD. (2007a). Ozone-induced 

symptoms on vegetation within the Moosehorn National Wildlife Refuge in Maine. Northeast Nat 14: 403-

414. http://dx.doi.org/10.1656/1092- 6194(2007)14[403:OSOVWT]2.0.CO;2 

Davis, DD. (2007a). Ozone-induced symptoms on vegetation within the Moosehorn National Wildlife Refuge in 3 

Maine. Northeast Nat 14: 403-414. http://dx.doi.org/10.1656/1092- 

6194(2007)14[403:OSOVWT]2.0.CO;2 

Davis, DD. (2007b). Ozone injury to plants within the Seney National Wildlife Refuge in northern Michigan. 

Northeast Nat 14: 415-424. 

Davis, DD. (2009). Ozone-induced stipple on plants in the Cape Romain National Wildlife Refuge, South Carolina. 

Southeastern Naturalist 8: 471-478. 

Davis, DD; Orendovici, T. (2006). Incidence of ozone symptoms on vegetation within a National Wildlife Refuge in 

New Jersey, USA. Environ Pollut 143: 555-564. http://dx.doi.org/10.1016/j.envpol.2005.10.051 
5-90 

Federal Register 1996. National Ambient Air Quality Standards for Ozone; Proposed Rule. 40 CFR 50; Federal 

Register 61: 65716 

Federal Register 1997. National Ambient Air Quality Standards for Ozone; Final Rule. 40 CFR 50; Federal Register 

62:38856 

Federal Register 2007. National Ambient Air Quality Standards for Ozone; Proposed Rule. 40 CFR 50; Federal 

Register 72: 37818 

Federal Register 2008. National Ambient Air Quality Standards for Ozone; Final Rule. 40 CFR parts 50 and 58; 

Federal Register 73:16436 

Federal Register 2010. National Ambient Air Quality Standards for Ozone; Proposed Rule. 40 CFR 50 and 58; 

Federal Register 75 FR: 2938 

Federal Register 2012. National Ambient Air Quality Standards for Oxides of Nitrogen and Sulfur; Final Rule. 40 

CFR 50; Federal Register 77 FR 20218 

Fiore, AM; Jacob, DJ; Field, BD; Streets, DG; Fernandes, SD; Jang, C. (2002). Linking ozone pollution and climate 

change: The case for controlling methane. Geophys Res Lett 29: 1919. 

http://dx.doi.org/10.1029/2002GL015601 

Frey, C. and Samet, J.M. (2012) CASAC Review of the EPA’s Policy Assessment for the Review of the Ozone 

National Ambient Air Quality Standards (First External Review Draft – August 2012). EPA-CASAC-13-

003. November 26, 2012. 

Frey, C. (2014) CASAC Review of the EPA’s Second Draft Policy Assessment for the Review of the Ozone National 

Ambient Air Quality Standards. EPA-CASAC-14-004. June 26, 2014. 

Grantz, DA; Gunn, S; Vu, HB. (2006). O3 impacts on plant development: A meta-analysis of root/shoot allocation 

and growth. Plant Cell Environ 29: 1193-1209. http://dx.doi.org/10.1111/j.1365-3040.2006.01521.x 

Gregg, JW; Jones, CG; Dawson, TE. (2003). Urbanization effects on tree growth in the vicinity of New York City 

[Letter]. Nature 424: 183-187. http://dx.doi.org/10.1038/nature01728 

Grulke, NE; Johnson, R; Esperanza, A; Jones, D; Nguyen, T; Posch, S; Tausz, M. (2003). Canopy transpiration of 

Jeffrey pine in mesic and xeric microsites: O3 uptake and injury response. Trees Struct Funct 17: 292-298. 

Grulke, NE; Minnich, RA; Paine, TD; Seybold, SJ; Chavez, DJ; Fenn, ME; Riggan, PJ; Dunn, A. (2008). 

Airpollution increases forest susceptibility to wildfires: A case study in the San Bernardino Mountains in 

southern California. In A Bytnerowicz; MJ Arbaugh; AR Riebau; C Anderson (Eds.), Wildland fires and 

air pollution; Section III: Ecological Impacts of Forest Fires and Air Pollution (pp. 365-403). Amsterdam, 

The Netherlands: Elsevier Ltd. http://dx.doi.org/10.1016/S1474-8177(08)00017-X 

Haefele, M., R.A. Kramer, and T.P. Holmes. (1991). Estimating the Total Value of a Forest Quality in HighElevation Spruce-Fir Forests. The Economic Value of Wilderness: Proceedings of the Conference. Gen. 

Tech. Rep. SE-78 (pp. 91-96). Southeastern For. Exper. Station. Asheville, NC: USDA Forest Service. 

Hall, Karen R. and Richard R. Braham. (1998). Native Pines of Eastern North America. North Carolina State 

University Department of Forestry. Available at http://www.ncsu.edu/project/dendrology/ 

Holmes, T., and R. Kramer. (1995). “An Independent Sample Test of Yea-Saying and StartingPoint Bias in 

Dichotomous-Choice Contingent Valuation.” Journal of Environmental Economics and Management 

28:121-132. 
5-91 

Heagle, AS. (1989). Ozone and crop yield*. Annu Rev Phytopathol 27: 397-423. 

http://dx.doi.org/10.1146/annurev.py.27.090189.002145

Heck, WW; Cowling, EB. (1997).The need for a long term cumulative secondary ozone standard - An ecological 

perspective. EM January: 23-33. 

Henderson, R. (2006) Letter from CASAC Chairman Rogene Henderson to EPA Administrator Stephen Johnson. 

October 24, 2006, EPA-CASAC-07-001. 

Henderson, R. (2008) Letter from CASAC Chairman Rogene Henderson to EPA Administrator Stephen Johnson. 

April 7,2008, EPA-CASAC-08-009. 

Hogsett, WE; Weber, JE; Tingey, D; Herstrom, A; Lee, EH; Laurence, JA. (1997). Environmental auditing: An 

approach for characterizing tropospheric ozone risk to forests. J Environ Manage 21: 105-120. 

http://dx.doi.org/10.1007/s002679900010

IPCC (Intergovernmental Panel on Climate Change). (2007). Climate change 2007: Impacts, adaptation and 

vulnerability. Cambridge, UK: Cambridge University Press. 

King, JS; Kubiske, ME; Pregitzer, KS; Hendrey, GR; McDonald, EP; Giardina, CP; Quinn, VS; Karnosky, 

DF.(2005). Tropospheric O3 compromises net primary production in young stands of trembling aspen, 

paper birch and sugar maple in response to elevated atmospheric CO2. New Phytol 168: 623-635. 

http://dx.doi.org/10.1111/j.1469-8137.2005.01557.x 

Kline, LJ; Davis, DD; Skelly, JM; Savage, JE; Ferdinand, J. (2008). Ozone sensitivity of 28 plant selections exposed 

to ozone under controlled conditions. Northeast Nat 15: 57-66. http://dx.doi.org/10.1656/1092- 

6194(2008)15[57:OSOPSE]2.0.CO;2 

Kohut, R. (2007). Assessing the risk of foliar injury from ozone on vegetation in parks in the US National Park 

Service's Vital Signs Network. Environ Pollut 149: 348-357. 

Kubiske, ME; Quinn, VS; Heilman, WE; McDonald, EP; Marquardt, PE; Teclaw, RM; Friend, AL; Karnoskey, DF. 

(2006). Interannual climatic variation mediates elevated CO2 and O3 effects on forest growth. Global 

Change Biol 12: 1054-1068. http://dx.doi.org/10.1111/j.1365-2486.2006.01152.x 

Kubiske, ME; Quinn, VS; Marquardt, PE; Karnosky, DF. (2007). Effects of elevated atmospheric CO2 and/or O3 on 

intra- and interspecific competitive ability of aspen. Plant Biol (Stuttg) 9: 342-355. 

http://dx.doi.org/10.1055/s-2006-924760 

Lee, EH; Hogsett, WE. (1996). Methodology for calculating inputs for ozone secondary standard benefits analysis: 

Part II. Research Triangle Park, NC: U.S. Environmental Protection Agency.

Lefohn, AS; Laurence, JA; Kohut, RJ. (1988). A comparison of indices that describe the relationship between 

exposure to ozone and reduction in the yield of agricultural crops. Atmos Environ 22: 1229-1240. 

http://dx.doi.org/10.1016/0004-6981(88)90353-8 

Lefohn, AS; Runeckles, VC. (1987). Establishing standards to protect vegetation - ozone exposure/dose 

considerations. Atmos Environ 21: 561-568. http://dx.doi.org/10.1016/0004-6981(87)90038-2 

Lefohn, AS; Jackson, W; Shadwick, DS; Knudsen, HP. (1997). Effect of surface ozone exposures on vegetation 

grown in the southern Appalachian Mountains: Identification of possible areas of concern. Atmos Environ 

31: 1695-1708. http://dx.doi.org/10.1016/S1352-2310(96)00258-0 
5-92 

Low, M; Herbinger, K; Nunn, AJ; Haberle, KH; Leuchner, M; Heerdt, C; Werner, H; Wipfler, P; Pretzsch, H; 

Tausz, M; Matyssek, R. (2006). Extraordinary drought of 2003 overrules ozone impact on adult beech trees 

(Fagus sylvatica). Trees Struct Funct 20: 539-548. http://dx.doi.org/10.1007/s00468-006-0069-z

Maier-Maercker, U. (1998). Predisposition of trees to drought stress by ozone. Tree Physiol 19: 71-78. 

McLaughlin, SB; Nosal, M; Wullschleger, SD; Sun, G. (2007a). Interactive effects of ozone and climate on tree 

growth and water use in a southern Appalachian forest in the USA. New Phytol 174: 109-124. 

http://dx.doi.org/10.1111/j.1469-8137.2007.02018.x 

McLaughlin, SB; Wullschleger, SD; Sun, G; Nosal, M. (2007b). Interactive effects of ozone and climate on water 

use, soil moisture content and streamflow in a southern Appalachian forest in the USA. New Phytol 174: 

125-136. http://dx.doi.org/10.1111/j.1469-8137.2007.01970.x 

Morgan, PB; Ainsworth, EA; Long, SP. (2003). How does elevated ozone impact soybean? Ameta-analysis of 

photosynthesis, growth and yield. Plant Cell Environ 26: 1317-1328.

Novak, K; Cherubini, P; Saurer, M; Fuhrer, J; Skelly, JM; Kräuchi, N; Schaub, M. (2007). Ozone air pollution 

effects on tree-ring growth, delta(13)C, visible foliar injury and leaf gas exchange in three ozone-sensitive 

woody plant species. Tree Physiol 27: 941-949.

National Park Service (NPS). (2006). Ozone Sensitive Plant Species in NPS Class I Areas. Available at 

http://www2.nature.nps.gov/air/Pubs/pdf/flag/NPSozonesensppFLAG06.pdf. 

NCSU College of Forest Resources (). Loblolly Pine. In Native Pines of Eastern North America. Available at 

http://www.ncsu.edu./project/dendrology/index/plantae/vascular/seedplants/gymnosperms/conifers/pine/pin

us/australes/loblollypine/html.

Pollastrini, M; Desotgiu, R; Cascio, C; Bussotti, F; Cherubini, P; Saurer, M; Gerosa, G; Marzuoli, R. (2010). 

Growth and physiological responses to ozone and mild drought stress of tree species with different 

ecological requirements. Trees Struct Funct 24: 695-704. http://dx.doi.org/10.1007/s00468-010-0439-4 

Richards, B.L., Middleton, J.T., and Hewitt, W.B. (1958). Air pollution with relation to agronomic crops. V. 

Oxidant stipple on grape. Agron. J. 50:559-561. 

Riikonen, J; Kets, K; Darbah, J; Oksanen, E; Sober, A; Vapaavuori, E; Kubiske, ME; Nelson, N; Karnosky, DF. 

(2008). Carbon gain and bud physiology in Populus tremuloides and Betula papyrifera grown under 

longterm exposure to elevated concentrations of CO2 and O3. Tree Physiol 28: 243-254. 

http://dx.doi.org/10.1093/treephys/28.2.243 Rose, A and Coulston, J.W. (2009). Ozone injury across the 

Southern United States, 2002 – 2006. Gen. Tech. Rep. SRS-118. Asheville, N.C.: US Department of 

Agriculture, Forest Service, Southern Research Station, pp. 25. 

Rose, A and Coulston, J.W. (2009). Ozone injury across the Southern United States, 2002 – 2006. Gen. Tech. Rep. 

18 SRS-118. Asheville, N.C.: US Department of Agriculture, Forest Service, Southern Research Station, 

pp. 19 25. 

Samet, J.M. (2010) Review of EPA’s Proposed Ozone National Ambient Air Quality Standard (Federal Register, 

Vol. 75, Nov. 11, January 19, 2010). EPA-CASAC-10-007. February 19, 2010. Available online at: 

http://yosemite.epa.gov/sab/sabproduct.nsf/264cb1227d55e02c85257402007446a4/610BB57CFAC8A41C

852576CF007076BD/$File/EPA-CASAC-10-007-unsigned.pdf 

Smith, G; Coulston, J; Jepsen, E; Prichard, T. (2003). A national ozone biomonitoring program: Results from field 

surveys of ozone sensitive plants in northeastern forests (1994-2000). Environ Monit Assess 87: 271-291. 
5-93 

Smith, G. (2012). Ambient ozone injury to forest plants in Northeast and North Central USA: 16 years of 

biomonitoring. Environ Monit Assess 184: 4049-4065. http://dx.doi.org/10.1007/s10661-011-2243-z 

Stevenson, DS; Dentener, FJ; Schultz, MG; Ellingsen, K; Van Noije, TPC; Wild, O; Zeng, G; Amann, M; Atherton, 

CS; Bell, N; Bergmann, DJ; Bey, I; Butler, T; Cofala, J; Collins, WJ; Derwent, RG; Doherty, RM; Drevet, 

J; Eskes, HJ; Fiore, AM; Gauss, M; Hauglustaine, DA; Horowitz, LW; Isaksen, ISA; Krol, MC; Lamarque, 

JF; Lawrence, MG; Montanaro, V; Muller, JF; Pitari, G; Prather, MJ; Pyle, JA; Rast, S; Rodriguez, JM; 

Sanderson, MG; Savage, NH; Shindell, DT; Strahan, SE; Sudo, K; Szopa, S. (2006). Multimodel ensemble 

simulations of present-day and near-future tropospheric ozone. J Geophys Res 111: D08301. 

http://dx.doi.org/10.1029/2005JD006338 

UNEP (United Nations Environment Programme). (2003). Millennium Ecosystem Assessment: Ecosystems and 

human well-being: A framework for assessment. Washington, DC: Island Press. 

University of Montana, College of Forest Conservation Wilderness Institute (2014). 

http://www.wilderness.net/printFactSheet.cfm?WID=583, accessed in 2014. U.M. College of Forestry and 

Conservation's Wilderness Institute in collaboration with the Arthur Carhart National Wilderness Training 

Center and Aldo Leopold Wilderness Research Institute.

USDA, Forest Service (2014) Tree basal area data, http://www.fs.fed.us/foresthealth/technology/nidrm2012.shtml

USDA, National Resources Conservation Service (2014), The PLANTS Database (http://plants.usda.gov, 2014), 

National Plant Data Center, Baton Rouge, LA.USDA Agricultural Research Service. (2012). Effects of 

Ozone Air Pollution on Plants. Available at http://www.ars.usda.gov/Main/docs.htm?docid=12462. 

USDA Economic Research Service. (2012). Background. Soil & Oilseed Crops. Available at 

http://www.ers.usda.gov/topics/crops/soybeans-oil-crops/background.aspx. 

USDA US Forest Service. (1990). Acer rubrum L. Red Maple. In Silvics of North America. Agriculture Handbook 

654, vol. 2 Hardwoods, Washington, DC. 877 p. Available at 

http://www.na.fs.fed.us/pubs/silvics_manual/volume_2/acer/rubrum.htm. 

USDA US Forest Service. (2014). Environmental Justice. Available at 

http://www.fs.fed.us/research/urban/environmental-justice.php. 

U.S. EPA (1989). Review of National Ambient Air Quality Standards for Ozone: Assessment of Scientific and 

Technical Information – OAQPS Staff Paper. Office of Air Quality Planning and Standards, Research 

Triangle Park, NC. 

U.S. EPA (1996). Air quality criteria for ozone and related photochemical oxidants [EPA Report]. (EPA/600/P93/004AF). U.S. Environmental Protection Agency, Research Triangle Park, NC. 

U.S. EPA (2006a). Air Quality Criteria for Ozone and Related Photochemical Oxidants (2006 Final). U.S. 

Environmental Protection Agency, Washington, DC. EPA/600/R-05/004aF-cF. March 2006. Available at: 

http://www.epa.gov/ttn/naaqs/standards/ozone/s_o3_cr_cd.html 

U.S. EPA (2006b). Ecological Benefits Assessment Strategic Plan. EPA-240-R-06-001. Office of Administrator, 

Washington, D.C. Available at http://yosemite.epa.gov/ee/epa/eed.nsf/webpages/ecologbenefitsplan.html 

U.S. EPA (U.S. Environmental Protection Agency). (2007). Review of the national ambient air quality standards for 

ozone: Policy assessment of scientific and technical information: OAQPS staff paper [EPA Report]. 

(EPA/452/R-07/003). Research Triangle Park, NC. 

http://www.epa.gov/ttn/naaqs/standards/ozone/data/2007_01_ozone_staff_paper
5-94 

U.S. EPA (2013). Integrated Science Assessment of Ozone and Related Photochemical Oxidants (Final). U.S. 

Environmental Protection Agency, Washington, DC. EPA/600/R-10/076F 

U.S. EPA (2014a). Welfare Risk and Exposure Assessment for Ozone. Office of Air Quality Planning and 

Standards, Research Triangle Park, NC, 27711. EPA-452/P-14-005a 

U.S. EPA. (2014b). Technology Transfer Network (TTN) Air Quality System (AQS). Available at 

http://www.epa.gov/ttn/airs/airsaqs. 

U.S. EPA. (2014c). Technology Transfer Network (TTN) Air Quality System (AQS). Download Detailed AQS 

Data. Available at http://www.epa.gov/ttn/airs/airsaqs/detaildata/downloadaqsdata.htm. 

U.S. EPA. (2014d). Air Trends: Design Values. Available at http://www.epa.gov/airtrends/values.html. 

US Forest Service, National Park Service, and U.S. Fish and Wildlife Service. 2010. Federal land managers’ air 

quality related values work group (FLAG): phase I report—revised (2010). Natural Resource Report 

NPS/NRPC/NRR—2010/232. National Park Service, Denver, Colorado. 

http://www.nature.nps.gov/air/Pubs/pdf/flag/FLAG_2010.pdfWest, JJ; Fiore, AM; Horowitz, LW; 

Mauzerall, DL. (2006). Global health benefits of mitigating ozone pollution with methane emission 

controls. PNAS 103: 3988-3993. http://dx.doi.org/10.1073/pnas.0600201103 

West, JJ; Fiore, AM; Naik, V; Horowitz, LW; Schwarzkopf, MD; Mauzerall, DL. (2007). Ozone air quality and 

radiative forcing consequences of changes in ozone precursor emissions. Geophys Res Lett 34: L06806. 

http://dx.doi.org/10.1029/2006GL029173

Wittig, VE; Ainsworth, EA; Long, SP. (2007). To what extent do current and projected increases in surface ozone 

affect photosynthesis and stomatal conductance of trees? A meta-analytic review of the last 3 decades of 

experiments [Review]. Plant Cell Environ 30: 1150-1162. http://dx.doi.org/10.1111/j.1365-

3040.2007.01717.x 

Wittig, VE; Ainsworth, EA; Naidu, SL; Karnosky, DF; Long, SP. (2009). Quantifying the impact of current and 

future tropospheric ozone on tree biomass, growth, physiology and biochemistry: A quantitative metaanalysis. Global Change Biol 15: 396-424. http://dx.doi.org/10.1111/j.1365-2486.2008.01774.x 

Young, T. F.; Sanzone, S., eds. (2002). A framework for assessing and reporting on ecological condition: an SAB 

report. Washington, DC: U.S. Environmental Protection Agency, Science Advisory Board; report no. 

EPASAB- EPEC-02-009. Available: http://www.epa.gov/sab/pdf/epec02009.pdf [9 December, 2003]. 
6-1 

6 CONSIDERATION OF ALTERNATIVE SECONDARY 

STANDARDS 

Chapter 5 reached the conclusion that the available evidence and exposure and risk 

information call into question the adequacy of public welfare protection provided by the current 

standard, and that it is appropriate for the Administrator to consider revising the current 

secondary standard to provide increased public welfare protection against O3-attributable effects 

on tree biomass loss, crop yield loss, and visible foliar injury, and their associated services, 

particularly for those effects associated with cumulative, seasonal exposures, to the extent these 

effects are judged adverse to the public welfare. Given that conclusion, this chapter describes 

the staff evaluation of the available body of evidence, and exposure, risk and air quality 

information with regard to support for consideration of alternative standards, as articulated by the 

following overarching question: 

 What alternative secondary standards are supported by the currently available 

scientific evidence, exposure/risk information and air quality analyses? 

To assist us in interpreting the currently available scientific evidence and the results of 

recent quantitative exposure/risk analyses to address this question, we have focused on a series 

of more specific questions in sections 6.1, 6.2 and 6.3 below. We consider both the scientific 

and technical information available at the time of the last review and information newly available 

since the last review which has been critically analyzed and characterized in the ISA. 

Specifically, we consider the currently available scientific evidence and technical information in 

the context of decisions regarding the basic elements of the NAAQS: indicator (section 6.1); 

averaging time and form (section 6.2); and level (section 6.3). CASAC advice on potential 

alternative standards is described in section 6.4 and staff conclusions on potential alternative 

standards are discussed in section 6.5. Section 6.6 summarizes staff conclusions on the adequacy 

of the current standard and the alternative standards that are appropriate for the Administrator to 

consider. Key uncertainties in this review and areas in which future research and data collection 

would better inform the next review are identified in section 6.7. 

6.1 INDICATOR 

With regard to the selection of an appropriate indicator for alternative secondary 

standards, we consider the following question. 
6-2 

 Does the information available in this review continue to support O3 as the 

indicator for ambient air photochemical oxidants? 

In the last review of the air quality for O3 and other photochemical oxidants and of the O3

standard, as in other prior reviews, the EPA focused on a standard for O3 as the most appropriate 

surrogate for ambient photochemical oxidants. Ozone is a long-established surrogate for 

ambient photochemical oxidants, among which it is by far the most widely studied with regard to 

effects on welfare and specifically on vegetation. The information available in this review adds 

to our understanding of the atmospheric chemistry for photochemical oxidants and O3 in 

particular (as described in the ISA, sections 3.2 and 3.6, and summarized in section 2.2 in this 

document). The 1996 Staff Paper noted that the database on vegetation effects is generally 

considered to raise concern at levels found in the ambient air for O3 and, therefore, control of 

ambient O3 levels has previously been concluded to provide the best means of controlling other 

photochemical oxidants of potential welfare concern (U.S. EPA, 1996b, p. 277). In the current 

review, while the complex atmospheric chemistry in which O3 plays a key role has been 

highlighted, no alternatives to O3 have been advanced as being a more appropriate surrogate for 

ambient photochemical oxidants. Ozone continues to be the only photochemical oxidant other 

than nitrogen dioxide that is routinely monitored and for which a comprehensive database exists 

(U.S. EPA, 2013, section 3.6). Thus, staff concludes that O3 remains the appropriate pollutant 

indicator for use in a secondary NAAQS that provides protection for public welfare from 

exposure to all photochemical oxidants. 

6.2 FORM AND AVERAGING TIME 

In considering potential forms and averaging times alternative to that of the current 

secondary standard (i.e., 4th highest daily maximum 8-hour average, averaged over 3 years), we 

address several specific questions. 

 To what extent does the currently available information provide support for 

considering forms different from that of the current secondary standard? 

In characterizing the current evidence, the ISA states that “[n]o recent information is 

available since 2006 that alters the basic conclusions put forth in the 2006 and 1996 O3 AQCDs” 

with regard to biologically relevant exposure indices (U.S. EPA, 2013, section 2.6.6.1, p. 2-43). 

Based on the current state of knowledge and the best available data assessed in this review, the 

ISA therefore concludes that exposure indices that cumulate and differentially weight the higher 

hourly average concentrations over a season and also include the mid-level values continue to 

offer the most scientifically defensible approach for use in developing response functions and in 

defining indices for vegetation protection. Quantifying exposures with indices that cumulate 

hourly O3 concentrations and preferentially weight the higher concentrations improves the 
6-3 

explanatory power of exposure/response models for growth and yield, over using indices based 

on mean and peak exposure values (U.S. EPA, 2013, section 2.6.6.1). These conclusions are 

based on the available body of evidence which provides a wealth of information, compiled over 

several decades, on the aspects of O3 exposure that are most important in influencing plant 

response. As discussed in the ISA, the importance of the duration of the exposure and the 

relatively greater importance of higher concentrations (over lower concentrations) in determining 

plant response to O3 have been well documented (U.S. EPA, 2013, section 9.5.3). Building on 

this research, other work has focused on developing “mathematical approaches for summarizing 

ambient air quality information in biologically meaningful forms for O3 vegetation effects 

assessment purposes ...” (U.S. EPA, 2013, section 9.5.2, p. 9-99), including those known as 

cumulative, concentration weighted forms (i.e., SUM06, W126). Much of this work was 

completed by the mid-1990s, and was summarized in the 1996 Criteria Document (CD) (U.S. 

EPA, 1996a, section 5.5). 

On the basis of this longstanding and extensive evidence demonstrating that the risk to 

vegetation comes from cumulative seasonal exposures, the EPA in the 1997 and 2008 reviews, as 

well as in the 2010 proposed rulemaking to reconsider the 2008 decision, recognized the 

importance of cumulative, seasonal exposures as a primary determinant of plant responses to O3

in ambient air (61 FR 65741-42; 62 FR 38878; 72 FR 37893, 37896, 37900, 37904; 73 FR 

16488-90, 16493-94; 75 FR 3000, 3010, 3012). For example, in the 1996 notice of proposed 

rulemaking, the Administrator recognized that the scientific evidence supported the conclusion 

that “a cumulative seasonal exposure index is more biologically relevant than a single event or 

mean index” (61 FR 65742). In the 2008 review, CASAC recognized that an important 

difference between the effects of short-term exposures to O3 on human health and the effects of 

O3 exposures on welfare is that “vegetation effects are more dependent on the cumulative 

exposure to, and uptake of, O3 over the course of the entire growing season” (Henderson, 2006, 

p. 5). In that review, the CASAC O3 Panel members were unanimous in supporting the final 

Staff Paper recommendation that “protection of managed agricultural crops and natural terrestrial 

ecosystems requires a secondary Ozone NAAQS that is substantially different from the primary 

ozone standard in averaging time, level and form” (Henderson, 2007, p. 3). Accordingly, in both 

the 1997 and 2008 reviews as well as the 2010 reconsideration, the Administrator proposed a 

secondary standard with a cumulative seasonal form as an appropriate policy option (61 FR 

65742-44; 72 FR 37899-905; 75 FR 3012-3027). 

In considering which exposure index was best suited for use as a form for the secondary 

O3 NAAQS, the 1996 CD and 1996 Staff Paper evaluated a variety of different types of forms. 

These documents noted that a number of forms (e.g., the one event, mean and unweighted 

cumulative SUM00) are unable to reliably predict plant response because they either ignore the 
6-4 

role of duration or ignore the disproportionate impact of higher concentrations by weighting all 

concentrations equally (U.S. EPA, 1996b, p. 224). Other forms that were considered at that time 

included multicomponent forms which take into account many other relevant factors (e.g., plant 

growth stage, predisposition from earlier exposures). Of all the different exposure forms, these 

multicomponent forms consistently predict plant response best. However, due to being speciesspecific and highly complex, they were not considered suitable for more general application in 

the context of standard setting (U.S. EPA, 1996b, pp. 224-225). On the other hand, 

concentration-weighted forms that take into account the role of duration and concentration 

perform almost as well as the multicomponent forms. These forms include several threshold 

forms (e.g., SUM06, AOT60) and sigmoidally weighted cumulative indices (e.g., W1261

) (U.S. 

EPA, 1996a, pp. 5-84 to 5-136; U.S. EPA, 1996b, pp. 223-227). Given that these cumulative 

concentration-weighted forms were able to similarly predict plant response on the datasets for 

which they were evaluated (e.g., NCLAN), it was not possible to distinguish between them on 

this basis. Partly as a result, CASAC deliberations in 1995 did not produce a consensus on 

which cumulative concentration-weighted form would be best suited for a secondary NAAQS. 

As discussed further in 6.3 below, a workshop held in January of 1996 provided a consensus 

recommendation on the SUM06 form as appropriate for use in secondary standards, while also 

recognizing that a W126 form could also be appropriate (Heck and Cowling, 1997). Subsequent 

to this, the final 1996 Staff Paper and 1996 proposal notice both identified the SUM06 form as 

appropriate to consider and propose, respectively (U.S. EPA, 1996b, p. 285, 61 FR 65716). In 

selecting the SUM06 form that imposed a threshold despite the lack of scientific evidence for a 

discernible threshold for O3-related vegetation effects across the range of studied species, the 

EPA noted that it had the benefit of not including concentrations that were considered at the time 

to be within the range of background, which was considered to be an important feature (U.S. 

EPA, 1996b, pp. 223-227). 

In the subsequent review, the form of the standard was revisited in light of continued 

evidence that there remained a lack of discernible threshold for vegetation effects in general, and 

newer estimates of O3 concentrations associated with background sources were lower than in the 

previous review such that their inclusion was less of a concern. On these bases, the 2007 Staff 

Paper recommended consideration of the W126 index as the basis for the form of a distinct 

secondary standard (U.S. EPA, 2006, pp. 9-11 to 9-15 and pp. AX9-159 to AX9-187; U.S. EPA, 

2007, pp. 7-15/16). The EPA then proposed two options for the secondary standard, one of 

 

1

 The W126 is a non-threshold approach described as the sigmoidally weighted sum of all hourly O3

concentrations observed during a specified diurnal and seasonal exposure period, where each hourly O3

concentration is given a weight that increases from 0 to 1 with increasing concentration (Lefohn et al, 1988; Lefohn 

and Runeckles, 1987; U.S. EPA, 2013, section 9.5.2). 
6-5 

which was to adopt a cumulative, seasonal standard based on the W126 index, while the other 

option was a secondary standard identical to the proposed revised primary standard (72 FR 

37818). The CASAC Panel in that review expressed preference for the W126 index (Henderson, 

2006). In deciding to reconsider the 2008 decision, the Administrator noted that past arguments 

or reasons for not moving to a cumulative, seasonal form, with appropriate exposure periods, 

were not based on disagreement over the biological relevance of the cumulative, seasonal form, 

or the recognized disadvantages of an 8-hour standard in measuring and identifying a specified 

cumulative, seasonal exposure pattern but were based on concerns over whether the EPA had an 

adequate basis to determine an appropriate level for a cumulative, seasonal secondary standard 

(75 FR 3019). Having reached the conclusion that such a level could be identified from within 

the range of levels proposed, the Agency proposed to set a secondary NAAQS in terms of a 

cumulative, seasonal standard form based on the W126 function (75 FR 2938). The CASAC 

also stated its support for this proposal, noting that it found the Agency’s reasoning to be 

supported by the extensive scientific evidence considered in the last review (Samet, 2010). 

In this review, we conclude that specific features associated with the W126 index still 

make it the most appropriate and biologically relevant cumulative concentration-weighted form 

for use in the context of the secondary O3 NAAQS review. In particular, the W126 index does 

not apply an arbitrary exposure threshold below which concentrations are not included. Given 

the acknowledged variability in vegetation sensitivity, including evidence that some species are 

sensitive at very low cumulative exposures, and the continued lack of evidence of an exposure 

threshold for effects above a W126 index of zero, such a feature is scientifically justifiable and 

desirable. Thus, we conclude that the W126 form is best matched to the evidence associated 

with vegetation effects, as well as addressing the policy-relevant issue of how to weight 

exposures associated with background sources. 

 To what extent does the currently available information provide support for 

consideration of a cumulative seasonal form derived as a sum of weighted O3

concentrations over daylight hours (8:00 am to 8:00 pm) and over the consecutive 

3-month period having the highest sum within the O3 season? 

 

As discussed in Chapter 5, mechanistic studies, including those recently assessed in this 

review, provide biological plausibility for the conclusions reached in the ISA that O3-induced 

effects on plants are cumulative, that higher concentrations appear to be more important in 

eliciting a response than lower concentrations, and that plant sensitivity to O3 can vary with time 

of day (U.S. EPA, 2013, p. 2-44). In particular, studies have shown that plants take up and 

respire gases through openings in their leaves, called stomata, which in general, are most open 

during daylight hours in order to allow sufficient CO2 uptake for use in carbohydrate production 
6-6 

through the light-driven process of photosynthesis. Ozone, when present in sufficient amounts, 

is taken up along with the CO2, where it and its derivatives can inhibit photosynthesis, leading to 

reduced carbohydrate production needed for growth, reproduction and repair (U.S. EPA, 2013, 

section 9.3.6; 75 FR 3013). Since plants are photosynthesizing during daylight hours and 

continue to grow throughout their growing season, the effects of repeated O3 exposures continue 

to accumulate, both on a diurnal and seasonal basis. Thus, for vegetation, the element of 

“averaging time” has more appropriately been considered in terms of relevant exposure periods – 

diurnal and seasonal -- over which exposures are cumulated, or summed. 

In the EPA’s consideration of such exposure periods in both the 1997 and 2008 reviews, 

and the 2010 reconsideration, the EPA identified the 12-hour daylight period from 8:00 am to 

8:00 pm as appropriately capturing the diurnal window with most relevance to the photosynthetic 

process (61 FR 35716; 72 FR 37900; 75 FR 2938, 3013). In so doing, the EPA recognized, as did 

CASAC, that in some parts of the country this period may not include all daytime hours or all 

exposures of importance to vegetation, thus potentially underestimating the impact of O3 at those 

sites (72 FR 37900-01; 75 FR 3013-14; Henderson, 2007, p. 3, pp. C-22-23). The evidence 

available in this review continues to provide support for focusing on the daylight hours, since for 

the majority of plants, the diurnal conditions of maximum O3 uptake occur mainly during the 

daytime hours (U.S. EPA, 2013, section 9.5.3.2). This evidence shows that, in general, (1) plants 

have the highest stomatal conductance during the daytime; (2) atmospheric turbulent mixing is 

greatest during the day in many areas; and (3) the high temperature and high light conditions that 

occur during the day and that typically promote the formation of tropospheric O3 also promote 

physiological activity in vegetation (U.S. EPA, 2013, section 9.5.3.2). 

In addition, as in past reviews, we have also considered the evidence available from a 

number of studies that have reported O3 uptake at night in some species (U.S. EPA, 2013, section 

9.5.3.2). Typically the rate of stomatal conductance at night is much lower than during the day. 

Across the studies discussed in the ISA, nocturnal conductance ranged from negligible to 25% of 

daytime values (U.S. EPA, 2013, section 9.5.3.2), and, in some studies, varied by season and 

drought conditions. However, many of these studies did not link the night-time flux to measured 

effects on plants, making it difficult to know in those studies whether the impacts on the plant 

from nocturnal exposures are greater or less than those from similar daytime exposures, and 

whether or not they should be considered as separate impacts or as additive or synergistic with 

impacts from the preceding or subsequent daytime exposure. 

Further, there are also uncertainties associated with the extent of the occurrence of high 

exposure to O3 at night. For significant nocturnal stomatal flux and O3 effects to occur, a 

susceptible plant with nocturnal stomatal conductance and low defenses must be growing in an 
6-7 

area with relatively high nighttime O3 concentrations (often high elevation sites) and appreciable 

nocturnal atmospheric turbulence. It is unclear how many areas there are in the U.S. where these 

atmospheric conditions occur. It may be possible that these conditions exist in mountainous 

areas of southern California, front-range of Colorado and the Great Smoky Mountains of North 

Carolina and Tennessee. However, more information is needed in locations with high nighttime 

O3 to assess the local O3 patterns, micrometeorology and responses of potentially vulnerable 

plant species (U.S. EPA, 2013, section 9.5.3.2). 

In consideration of the uncertainties that remain regarding the importance and extent of 

nocturnal exposures associated with plant uptake, and whether and how they might be 

incorporated into a national index, we conclude that it is appropriate to continue to focus on the 

12-hour daylight exposure period of 8:00 am to 8:00 pm. We note that available monitoring data 

indicates that the daily increase in O3 concentrations generally does not begin until after 8:00 am 

(U.S. EPA, 2013, section 3.6.3.2). In regard to this staff conclusion on an appropriate diurnal 

exposure period, CASAC states that “[a]ccumulation over the 08:00 a.m. – 08:00 p. m. daytime 

12-hour period is a scientifically acceptable and recommended means of generalizing across 

latitudes and seasons” (Frey, 2014a, p. 13). 

With regard to identification of the seasonal period over which to cumulate exposures, we 

note that a plant is vulnerable to O3 pollution as long as it has foliage and is physiologically 

active (U.S. EPA, 2013, section 9.5.3, p. 9-112), i.e., during its growing season. The length of 

vegetative growing seasons varies depending on the type or species of vegetation and where it 

grows. For example, as discussed in the ISA, annual crops are typically grown for periods of 

two to three months while perennial species may be photosynthetically active longer, up to 12 

months each year for some species (U.S. EPA 2013, section 9.5.3, p. 9-112). In general, the 

period of maximum physiological activity and thus potential O3 uptake for vegetation coincides 

with some or all of the intra-annual period defined as the O3 season, which can vary on a stateby-state basis (U.S. EPA, 2013, Figure 3-24, p. 3-83). This is because the high temperature and 

high light conditions, which can vary geographically, typically promote the formation of 

tropospheric O3, as well as physiological activity in vegetation (U.S. EPA, 2013, section 9.5.3, p. 

9-112). 

The exposure periods used in studies of O3 effects on vegetation reflect this 

understanding, with crop studies typically using shorter seasonal exposure periods and studies of 

longer lived trees and other perennial vegetation often extending for the entire annual growing 

season or in some cases over multiple growing seasons. Specifically, the ISA notes that “[m]ost 

of the crop studies done through NCLAN had exposures less than three months with an average 

of 77 days. Open-top chamber studies of tree seedlings, compiled by the EPA, had an average 

exposure of just over three months or 99 days. In more recent FACE experiments, SoyFACE 
6-8 

exposed soybeans for an average of approximately 120 days per year and the Aspen FACE 

experiment exposed trees to an average of approximately 145 days per year of elevated O3, 

which included the entire growing season at those particular sites” (U.S, EPA, 2013, section 

9.5.3.2, p. 9-112). Further, the U.S. Forest Service and federal land managers have typically 

used the 6 months from April through September as the accumulation period (U.S, EPA, 2013, 

section 9.5.3.2, p. 9-112). However, despite the possibility that plants may be exposed to 

ambient O3 longer than 3 months in some locations, the ISA notes that “[t]he exposure period in 

the vast majority of O3 exposure studies conducted in the U.S. has been much shorter than 6 

months...” and “there is generally a lack of exposure experiments conducted for longer than 3 

months” (U.S. EPA, 2013, section 9.5.3.2, p. 9-112). As a result, analyses of effects in terms of 

the W126 exposure index have typically defined the index in terms of a 3-month exposure period 

or at least in terms of periods shorter than 6 months (e.g., SoyFACE, Aspen FACE) (U.S, EPA, 

2013, p. 9-112). 

In the current review, the EPA conducted a new analysis to further inform the 

consideration of the most appropriate seasonal accumulation period (U.S. EPA, 2013, section 

9.5.3). This analysis calculated and compared the 3- and 6-month maximum W126 index values 

for over 1,200 AQS and CASTNET EPA monitoring sites for the years 2008-2009. The two 

accumulation periods were found to be highly correlated metrics (U.S. EPA, 2013, Figure 9-13; 

section 9.5.3). The analysis indicates that in the U.S., W126 cumulated over 3 months and W126 

cumulated over 6 months could be proxies of one another, as long as the period in which daily 

W126 is accumulated corresponds to the seasonal maximum. Therefore, it is expected that either 

statistic will predict vegetation response equally well. Given the above information, and in 

particular the results of the EPA analysis showing the maximum 3-month period is highly 

correlated with the longer 6-month maximum period, we again conclude that it is appropriate to 

continue to focus on the consecutive 3-month period with the highest cumulative exposure value 

within the monitored O3 season as the seasonal exposure period with most relevance to 

vegetation. Given its review of the available science, CASAC also expressed support for this 

seasonal period, stating that “[t]he Second Draft PA makes a very strong case, consistent with 

previous CASAC judgment, for changing the secondary metric to the W126 averaged over the 

highest three-month interval” (Frey, 2014a, p. 13). 

6.3 LEVEL 

In considering potential levels for alternative secondary standards, we again find it useful 

to note that the protection provided by the secondary standard derives from the combination of 

all elements of the standard (indicator, form, averaging time, and level). Thus, in light of the 

discussions in section 6.2 above, we first consider what level or range of levels can reasonably be 
6-9 

judged to provide a requisite degree of public welfare protection when combined with a W126 

index form of cumulatively weighted concentrations from 8:00 am to 8:00 pm over a maximum 

consecutive 3-month period. In addition to considering the information in the context of a single 

growing season, we also consider it in the context of a form for this W126 metric averaged 

across three consecutive growing seasons for reasons discussed in section 6.2 above. 

 In the discussion below, we turn first to consideration of the currently available 

scientific evidence as assessed and characterized in the ISA. We then consider the WREA 

findings with regard to vegetation, ecosystem effects and services estimated for different air 

quality scenarios. We additionally take note of important uncertainties and limitations in the 

evidence and exposure/risk analyses, as well as considerations related to interpreting these 

impacts in light of the additional policy considerations described in the adversity paradigm. 

Lastly, we take note of judgments to be made by the Administrator in drawing conclusions 

regarding effects and risks that represent adverse effects to public welfare. In so doing, we 

identify key considerations with regard to the currently available evidence, exposure/risk 

information and associated uncertainties in identifying the range of levels that may be 

appropriate to consider for a cumulative seasonal secondary standard. Such levels are described 

in section 6.5 below, which describes staff conclusions regarding alternative secondary standards 

appropriate to consider in this review. 

 What does the currently available evidence indicate with regard to the range of 

W126-based index values that may provide protection from vegetation effects of 

O3? 

In answering this question, we first consider quantitative evidence for O3 exposure effects 

on plant growth, productivity and related endpoints. In so doing, we draw primarily on the 

robust E-R functions developed in OTC studies for tree seedling and crop species as described in 

the ISA (U.S. EPA 2013, section 9.6), and as used in the WREA exposure and risk analyses 

(U.S. EPA, 2014b, section 6.2), and discussed in Chapter 5 of this document (Figures 5-1 and 5-

4). It is important to note that these functions are used to provide estimates of growth and yield 

reduction in tree seedlings and crops that might be expected to result from exposure over a single 

growing season to various O3 concentrations expressed in terms of a W126 index. We also 

consider the available evidence and exposure/risk information for visible foliar injury. 

As a point of clarification, we note that CASAC commented that it “concurs that relative 

biomass loss for tree species, crop yield loss, and visible foliar injury are appropriate surrogates 

of a wide range of damage that is adverse to public welfare.” (Frey, 2014a, p. 10). While we 

agree that effects at the individual tree, crop, or other plant species level, in and of themselves, 

can be directly related to effects on the public welfare when they occur to a sufficient degree on 

lands with an intended use that can be affected by O3-induced vegetation effects (e.g., timber 
6-10 

production, AQRVs in Class I areas), we also caution that not all predicted effects on vegetation 

occur to such a degree or occur on lands within this category. Thus, in considering the predicted 

effects on studied tree and crop species under various W126 exposures, we are mindful of the 

need to further determine under what conditions they can be considered surrogates for impacts 

that are important in the public welfare policy context. 

 Table 6-1 below presents estimates of relative biomass and yield loss for the 11 and 10 

studied species of tree seedlings and crops, respectively, for which we have robust E-R functions 

developed in OTC studies, for a single growing season exposure to a number of W126 index 

values. In this table, we have included observations related to median and individual species 

relative biomass loss in tree seedlings and relative crop yield loss, at the target benchmark levels 

of 2% and 5%, respectively. These benchmarks are consistent with the 2% and 5% benchmarks 

for tree seedlings and crops, respectively, as advised by CASAC in this review (Frey, 2014a; 

section 6.4 below), and with values given focus in the 1996 expert consensus workshop. We 

have also included information on the number of studied species with estimates below other 

benchmarks that may also be of interest (i.e., 5%, 10%, and 15% for trees and 10% for crops). 

CASAC has placed most emphasis on the median species response in recommending a range of 

scientifically supportable levels. For example, CASAC noted that “[i]n our scientific judgment, it 

is appropriate to identify a range of levels of alternative W126-based standards that includes 

levels that aim for not greater than 2% RBL for the median tree species” (Frey, 2014a, p. 14).” 

CASAC also recognizes that as a policy matter the Administrator may find it useful to also 

consider information related to individual species responses, to the degree that they have special 

significance to the public welfare, when selecting an appropriate level or range of levels. 

Specifically, CASAC states that “[a]s a policy recommendation, separate from its advice above 
6-11 

Table 6-1. Tree seedling biomass loss and crop yield loss estimated for O3 exposure over a 

season. 

W126 value 

for exposure 

period

Tree seedling biomass lossA Crop yield lossC

Median Value Individual Species Median Value Individual Species 

21 ppm-hrs Median species w. 

6.8% loss B

< 2% loss: 3/11 species 

< 5% loss: 5/11 species 

<10% loss: 7/11 species 

<15% loss: 10/11 species 

>40% loss: 1/11 species 

Median species w. 

7.7 % loss D

< 5% loss: 4/10 species 

>5,<10% loss: 3/10 species 

>10,<20% loss: 3/10 species 

19 ppm-hrs Median species w. 

6.0% loss B

< 2% loss: 3/11 species

<5% loss: 5/11 species 

<10% loss: 7/11 species 

<15% loss: 10/11 species 

>30% loss: 1/11 species 

Median species w. 

6.4 % loss D

< 5% loss: 5/10 species 

>5, <10% loss: 3/10 species 

>10,<20% loss: 2/10 species 

17 ppm-hrs Median species w. 

5.3% loss B

< 2% loss: 5/11 species

<5% loss: 5/11 species 

<10% loss: 9/11 species 

<15% loss: 10/11 species 

>30% loss: 1/11 species 

Median species w. 

5.1 % loss D

< 5% loss: 5/10 species 

>5, <10% loss: 3/10 species 

>10,<20% loss: 2/10 species 

15 ppm-hrs Median species w. 

4.5% loss B

< 2% loss: 5/11 species 

<5% loss: 6/11 species 

<10% loss: 10/11 species 

>30% loss: 1/11 species 

Median species 

w.<5% loss D

< 5% loss: 6/10 species 

>5, <10% loss: 4/10 species 

13 ppm-hrs Median species w. 

3.8% loss B

< 2% loss: 5/11 species 

<5% loss: 7/11 species 

<10% loss: 10/11 species 

>20% loss: 1/11 species 

Median species 

w.<5% loss D

< 5% loss: 6/10 species 

>5, <10% loss: 4/10 species 

11 ppm-hrs Median species w. 

3.1% loss B

< 2% loss: 5/11 species 

<5% loss: 8/11 species 

<10% loss: 10/11 species 

>20% loss: 1/11 species 

Median species 

w. <5% loss D

< 5% loss: 9/10 species 

>5, <10% loss: 1/10 species 

9 ppm-hrs Median species w. 

2.4% loss B

< 2% loss: 5/11 species 

<5% loss: 10/11 species 

>20% loss: 1/11 species 

Median species 

w. <5% loss D

< 5% loss: all species 

7 ppm-hrs Median species w. 

<2% loss B

< 2% loss: 7/11 species 

<5% loss: 10/11 species 

>15% loss: 1/11 species 

Median species 

w. <5% loss D

< 5% loss: all species 

A Estimates here are based on the 11 E-R functions for tree seedlings described in WREA, Appendix 6F and discussed in section 

5.2.1, with the exclusion of cottonwood. See CASAC comments (Frey, 2014a). 

B This median value is the median of the composite E-R functions for 11 tree species in the WREA, Appendix 6F (also discussed 

in section 5.2.1). 

C Estimates here are based on the 10 E-R functions for crops described in Appendix 6F and discussed in section 5.3.1. 

D This median value is the median of the composite E-R functions for 10 crops from WREA, Appendix 6F (also discussed in 

section 5.3.1). 
6-12 

regarding scientific findings, the CASAC advises that a level of 15 ppm-hrs for the highest 3-

month sum in a single year is requisite to protect crop yield loss, but that lower levels provide 

additional protection against crop yield loss. Furthermore, there are specific economically 

significant crops, such as soybeans, that may not be protected at 15 ppm-hrs but would be 

protected at lower levels” (Frey, 2014a, p. iii). 

From Table 6-1, we see that median tree species biomass loss is at or below 2% only at 

the lowest W126 index value assessed, 7 ppm-hrs. As the W126 index value is incrementally 

increased, median RBL also increases incrementally, so that at W126 index values of 9, 11, 13, 

15, and 17, the median RBL increases to 2.4%, 3.1%, 3.8%, 4.5%, and 5.3%, respectively. Thus 

over the W126 range of 7 to 17 ppm-hrs, median species biomass loss ranges from 

approximately 2% to approximately 5%. 

We also believe it is informative to examine the individual species responses and RBL 

over the same W126 range (7 to 17 ppm-hrs). We first note, based on Figure 5-1 (B) in chapter 

5, that over this range, five species maintain RBLs of less than 2%. These more tolerant species 

include Douglas fir, loblolly pine, Virginia pine, sugar maple and red maple. Thus, little 

additional protection would be achieved for these species below the W126 index value of 17 

ppm-hrs. Two of these species (red maple and sugar maple) would only exceed 2% RBL at 21 

ppm-hrs. In contrast, black cherry, the most sensitive of the remaining six species, has RBL 

ranging from approximately 36% at 17 ppm-hrs down to approximately 17% at 7 ppm-hrs. 

Thus, given that the magnitude of predicted black cherry RBL could be judged adverse over this 

range, it is not clear to what extent this information informs the selection of an appropriate level 

(Table 6-1; U.S. EPA, 2014b, section 6.2, Appendix 6A), though clearly protection would be 

expected to be greater at lower W126 index values. We further note that CASAC based their 

recommendation of an appropriate W126 level by considering the median tree species RBL of no 

more than 2%, but some levels within CASAC’s recommended range allow for the possibility 

for individual species RBL to go much higher. 

Because Table 6-1 was updated in this final PA to deemphasize cottonwood, based on 

staff’s understanding of CASAC advice in that regard, we note that the CASAC advice based on 

the numbers of species protected to no more than 2% RBL and median RBL values for tree 

seedlings associated with various W126 levels, as shown in the in the second draft PA table 

(U.S. EPA, 2014a), is no longer consistent in some cases with the revised Table 6-1. For 

example, in commenting on the version of Table 6-1 in the second draft PA CASAC states that 

“[t]able 6-1 presents the RBL results for individual species for different levels of W126. This 

table demonstrates that a range of 7 ppm-hrs to 15 ppm-hrs will protect against RBL of 2% for at 
6-13 

least 5 of the 12 species”2

 (Frey, 2014a, p. 14). In addition, CASAC states that “[t]he CASAC 

does not support a level higher than 15 ppm-hrs. For example, at 17 ppm-hrs, the median tree 

species has 6% relative biomass loss.... These levels are unacceptably high” (Frey, 2014a, p. iii). 

While we continue to place weight on CASAC’s scientific judgments that a 6% median RBL is 

unacceptably high, and that the 2% median RBL is an important benchmark to consider, we also 

note that the updated median RBL for a W126 level of 17 ppm-hrs is now 5.3%. 

We further note that CASAC does not provide additional clarification regarding its views 

on the acceptability of median tree species RBL levels between 2% and 6%, beyond noting that 

values closer to the lower end of the range (W126 index value of 7 ppm-hrs) would provide 

greater protection for more sensitive tree species, and that the levels within CASAC’s 

recommended range allow for the possibility for individual species RBL to go much higher than 

2% and 6%. Given the nature of this input, we then considered the RBL information available for 

the remaining five species (i.e., eastern white pine, aspen, tulip poplar, ponderosa pine, red alder) 

to further inform our evaluation of the additional protection that potentially could be achieved at 

different W126 levels within the range being considered. We thus note that at the W126 index 

value of 17 ppm-hrs, one species (eastern white pine) has RBL above 10% and one species (red 

alder) has RBL of 5.3% (below 6%) while the other three fall between approximately 6.7% and 

9.8%. At the W126 index value of 15 ppm-hrs, two (i.e., red alder and ponderosa pine) of the 

five species fall below 6% RBL, while the remaining 3 species have RBLs that range from 7.4% 

to 8.8%. At the W126 index value of 13 ppm-hrs, three species (i.e., tulip poplar, ponderosa 

pine, red alder) fall below 6% RBL, while the remaining two have RBLs of 7.0% and 7.1%. At 

the W126 index value of 11 ppm-hrs, all five species have RBLs below 6%. Taken together with 

the more tolerant species, the proportion of the studied tree species with RBLs below 6% are 

6/11, 7/11, 8/11, and 10/11 at W126 index values of 17, 15, 13, and 11 ppm-hrs, respectively. 

To the extent the focus is placed on different % RBL benchmarks and the proportion of studied 

trees protected at those levels, as well the expected impacts to associated ecosystem services, in 

regard to the identification of the appropriate level or range of levels, this information may be 

appropriate to consider. 

With respect to crops, based on the 10 robust E-R functions (i.e., barley, lettuce, field 

corn, grain sorghum, peanut, winter wheat, cotton, soybean, potato and kidney bean) described in 

the ISA and additionally analyzed in the WREA (Figure 5-4), Table 6-1 shows that for the 

CASAC recommended target benchmark protection level of 5% for median crop relative yield 

loss (RYL), W126 index values ranging from 7 to 17 ppm-hrs are protective. However, when 

 

2

 We note that the updated table shows that a range of 7 ppm-hrs to 17 ppm-hrs will protect against RBL of 

2% for at least 5 of the 11 species. 
6-14 

individual species are considered over this same range, the proportion of crops protected varies 

from 5/10, 6/10, 7/10, 9/10, 10/10, and 10/10 at the W126 levels of 17, 15, 13, 11, 9, and 7 ppmhrs. To the extent a given species is judged as having particular importance to the public 

welfare, breaking the information down by species can be helpful. For example, less than 5% 

yield loss was estimated for soybeans at the W126 index value of 12 ppm-hrs (U.S. EPA 2014, 

Figure 6-3). Four of the studied crop species (barley, lettuce, field corn, and grain sorghum) are 

more tolerant, with RYL under 1% over the W126 range from 7 to 17 ppm-hrs. Peanut also 

remained under 4% RYL over the same W126 range. Other species differed regarding the W126 

level at which RYL reached or fell below 5%. Specifically, for winter wheat, cotton, soybean, 

kidney bean and potato, the relevant W126 index values at which RYLs were below 5% are 15, 

13, 11, 11, and 9 ppm-hrs. As noted in Chapter 5, the significance of these predicted RYLs to 

the public welfare could be informed by the recognition that crops are heavily managed to obtain 

the desired yield, and the extent to which yield reductions in any specific crop in a particular 

location are considered adverse to public welfare could depend on a number of economic factors, 

including crop prices, crop substitution, and the welfare importance of relative changes in 

consumer and producer surplus.

With respect to considering visible foliar injury, and its ability to inform selection of an 

appropriate target level or range of levels for public welfare protection, we first recognize its 

value as a long-standing and well-established bioindicator of O3 exposure, as described in the 

ISA (U.S. EPA 2013, section 9.4.2). In addition to the role of visible foliar injury as an 

indicator, we note that the aesthetic aspects of visible foliar injury itself have the potential to be 

important to public welfare (as described in section 5.4). CASAC also “concurs that visible 

foliar injury can impact public welfare by damaging or impairing the intended use or service of a 

resource. Visible foliar injury that is adverse to public welfare can include: visible damage to 

ornamental or leafy crops that affects their economic value, yield, or usability; visible damage to 

plants with special cultural significance; and visible damage to species occurring in natural 

settings valued for scenic beauty or recreational appeal” (Frey, 2014a, p. 10). In this regard, we 

first note that several tribes have identified a number of O3 sensitive species that are important to 

their cultural practices (Appendix 5A). These species have many cultural uses such as food, 

medicines, dyes, tools/textiles, spiritual, and commercial. In addition, visible foliar injury has 

been identified by the federal land managers (FLMs) as a diagnostic tool for informing 

conclusions regarding potential ozone impacts on potentially sensitive AQRVs (USFS, NPS, 

FWS, 2010), and the evidence shows that injury has been documented in such areas under recent 

air quality conditions. 
6-15 

Despite its recognized importance to the public welfare as a general matter, we are 

unaware of any injury benchmarks or criteria that have been identified by the FLMs as to what 

extent and/or severity of observed foliar injury warrants protection efforts. In considering 

CASAC comments in this regard we note that while it states that “[a] level below 10 ppm-hrs is 

required to reduce foliar injury” (Frey, 2014a, p. iii), CASAC does not provide any additional 

information regarding the public welfare significance of this degree of injury or what an 

appropriate target benchmark or range of benchmarks would be for foliar injury in relation to 

what could be considered adverse to the public welfare. Given that there is substantial variability 

in this endpoint, such that “the degree and extent of visible foliar injury development varies from 

year to year and site to site ... even among co-members of a population exposed to similar O3

levels, due to the influence of co-occurring environmental and genetic factors” (U.S. EPA, 2013, 

p. 9-38), staff recognizes the lack of a consistent or generally predictable relationship between 

particular W126 exposures and visible foliar injury incidence. We additionally note uncertainty 

in what can be concluded from foliar injury in relation to plant health, productivity and 

ecological function as “it is not presently possible to determine, with consistency across species 

and environments, what degree of injury at the leaf level has significance to the vigor of the 

whole plant” (U.S. EPA, 2013, p. 9-39). However, we do recognize the Congressional mandate, 

provided in the CAA amendments of 1977 that establish additional protections for Class I areas. 

The 1997 Consensus Workshop Report (Heck and Cowling, 1997) discussed below, noted the 

potential for visible foliar injury to occur at very low levels. CASAC also stated that “[v]isible 

foliar injury is even more sensitive than RBL of 2%, with W126 index values below 10 ppm-hrs 

required to reduce the number of sites showing visible foliar symptoms” (Frey, 2014, p. 14).We 

further note that the information discussed here regarding incidence of visible foliar injury does 

not include information regarding the severity of the observed symptoms and the degree to which 

the public welfare impacts from different severity benchmarks might vary. Thus, there is 

additional uncertainty regarding the potential variability in the severity of the symptoms across 

species and locations and to what degree this would affect the public welfare significance of 

these effects so that the appropriate range of W126 index values to protect against this effect is 

difficult and complicated to identify. 

In further considering the available information pertaining to the question above, we 

additionally recognize conclusions that have been drawn by expert committees with regard to 

these endpoints (i.e., tree seedling growth, crop yields and visible foliar injury). For example, in 

their review of staff documents during the O3 NAAQS review completed in 1997, the CASAC 

O3 panel members expressed a wide range of opinions on aspects of the evidence important to 

consider in judging the adequacy of the O3 secondary standard and in considering the form and 

level that would be appropriate for a secondary O3 standard (Wolff, 1996). Subsequent to 
6-16 

CASAC meetings in 1995 on this topic, a consensus-building workshop sponsored by the 

Southern Oxidant Study group was held on the topic of the O3 secondary standard in January 

1996 (Heck and Cowling, 1997). This workshop was attended by 16 scientists with backgrounds 

in agriculture, managed forest, natural systems, and air quality, all of whom were leaders in their 

fields and whose research formed the basis of much of the research examined in the 1996 

Criteria Document. These scientists expressed their judgments on what standard level(s) would 

provide vegetation with protection from O3-related adverse effects that would be adequate, in 

their view.3,4 As the 1997 workshop publication indicates, the scientists at the 1996 workshop 

also reached consensus views regarding the types of exposures that were important in eliciting 

plant response and the types of metrics that were best at predicting these responses (Heck and 

Cowling, 1997). Before coming to agreement on daily and seasonal durations and forms 

pertinent to a distinct secondary standard, the participants discussed and identified endpoints to 

consider for natural, forest and agricultural ecosystems.5 With regard to form of the standard, 

participants concurred with either the SUM06 or W126 metrics, with consensus finally reached 

for SUM06, with some qualification regarding implications for a threshold. The participants 

identified the ranges they felt should be considered for each of three endpoints. Overall, the 

SUM06 values ranged from 8 to 20 ppm-hrs corresponding to W126 index values ranging from 5 

to 17 ppm-hrs, based on the EPA analysis focused on conditions in NCLAN studies.6 This 

overall range reflected ranges for each of the three endpoints, with the following considerations 

(Heck and Cowling, 1997). 

– Crops (yield reductions): SUM06 of 15-20 ppm-hrs (13 to 17 ppm-hrs, W126). 

This range was recognized to generally consider <10% yield loss in more than 

75% of species. 

– Trees (growth effects): SUM06 of 10-16 ppm-hrs (7 to 14 ppm-hrs, W126). This 

range was recognized to generally consider 1-2% per year growth reduction; in so 

 

3

 At the time of the workshop, the secondary O3 standard being reviewed by EPA was a 1-hour average of 

0.12 ppm (identical to the primary standard at that time). In 1997, EPA concluded the review by revising both 

standards to a longer averaging time of 8 hours with a level of 0.08 ppm (62 FR 38856). 4

 The workshop publication describes the primary objective for the workshop as having been to assemble 

knowledgeable scientists to develop a group consensus on “various critical components associated with a possible 

revised secondary ozone standard” (Heck and Cowling, 1997). 5

 For natural ecosystems, they focused on foliar injury as an indicator. For forest ecosystems, they 

concluded the data did not support selection of an indicator of effects on forest structure or function. As a result, 

they identified two indicators pertinent to the systems: growth effects on seedlings from species of natural forest 

stands (1-2% per year reduction), and growth effects on seedlings and saplings from tree plantations (1-2% per year 

reduction). For agricultural systems, the participants focused on protection against crop yield reductions, with their 

acknowledgment of high uncertainties at 5% leading them to a crop yield endpoint of 10% yield reduction (Heck 

and Cowling, 1997). 6

During the last review, W126 index values corresponding to the SUM06 values cited in the report were 

estimated using the NCLAN crop loss data, a key dataset considered by workshop participants (see Appendix 7B of 

2007 Staff Paper; Appendix 6A of this document). 
6-17 

doing, the group identified a need to consider the potential for year-to-year 

compounding of impacts in long-lived perennial species. 

– Visible Foliar Injury: SUM06 of 8 to 12 ppm-hrs (5 to 9 ppm-hrs, W126). 

Since the publication of 1996 workshop report and conclusion of the 1997 NAAQS 

review, the evidence base has continued to expand as described in the 2006 CD and ISA (U.S. 

EPA, 2006; U.S. EPA, 2013). With regard to tree growth effects and crop yield reductions, 

results of additional studies conducted in the field have confirmed the tree seedling biomass loss 

and crop yield loss E-R relationships derived from earlier studies that used OTC (U.S. EPA 

2013, section 9.6). 

In the 2008 review, CASAC provided comments related to a cumulative seasonal 

secondary standard in the context of their comments on the draft and final Staff Papers and on 

the final decision (Henderson, 2006; Henderson, 2007; Henderson, 2008). In all instances, they 

conveyed support for establishment of a distinct secondary standard with a cumulative seasonal 

form. While the EPA, in the 2007 Staff Paper and 2007 notice of proposed rulemaking, 

recognized a broader range of W126 index values as appropriate for consideration with regard to 

a distinct secondary standard, the CASAC Panel focused on a range they described as 

approximately equivalent to that identified by the 1996 workshop participants (Henderson, 2007, 

pp. 3, C-27).7

 In the CASAC Panel 2006-2007 advice on levels for such a standard, their 

suggestion was a focus on levels for a W126 index approximately equivalent to a SUM06 range 

of 10 to 20 ppm-hrs (Henderson, 2006, 2007, 2008), which they estimated in 2007 to be a range 

from 7 (or 7.5) to 15 ppm-hrs. Based on their consideration of the information available in that 

review (with regard to potential magnitude of effects across multiple years), the CASAC Panel 

further advised that “[i]f multi-year averaging is employed to increase the stability of the 

secondary standard, the level of the standard should be revised downward to assure that the 

desired threshold is not exceeded in individual years” (Henderson, 2007, p. 3). The CASAC 

advice provided on the 2010 proposed reconsideration and in this review is summarized in 

section 6.4 below. 

In considering the evidence briefly summarized above in the context of levels for a 

W126-based standard, we recognize that given the different types of O3-induced effects, genetic 

variability within and between species, and environmental modifiers of effects that also 

contribute to variability, it is not feasible to identify a range of cumulative seasonal exposures 

from the vegetation effects evidence which would provide a consistent degree of protection for 

 

7

 Appendix C of the March 26, 2007 CASAC letter (Henderson, 2007) used a 2001 ambient concentration 

dataset and other factors, rather than study data considered in the 1996 workshop, in estimating an “equivalency” 

between the two indices. 
6-18 

all species. Thus, in our consideration of the evidence, we note the importance of considering 

several dimensions that pertain to judgments regarding public welfare significance. For 

example, we take note of the usefulness of considering the cumulative seasonal exposure at 

which the median species response or the majority of the species’ responses are expected to be 

below minimal response benchmarks of interest and at which only a very few species’ responses 

are expected to exceed more substantial response benchmarks. Before articulating such 

considerations with regard to specific benchmarks and index values, we first consider the WREA 

findings in the context of the following question. 

 What are the nature and magnitude of risks to vegetation estimated for the 

average W126 index scenarios evaluated in the WREA, and what is the 

magnitude of risk reduction from risks estimated for air quality conditions 

estimated for the current standard? 

The WREA provides a characterization of ambient O3 exposure and its relationship to 

ecological effects, and estimates of the resulting impacts to several ecosystem services. In 

considering the question posed above, we focus particularly on WREA estimates related to O3

effects on plant biomass and associated ecosystem services effects. The WREA analyses provide 

information on the geographical extent of the effects of O3 exposure on plant biomass for 

different air quality scenarios. We also note the relationships among effects on individual plants 

to other ecosystem components and functions, such as carbon sequestration and air pollutant 

removal (U.S. EPA 2013, section 9.4.3.4; U.S. EPA, 2014b, sections 6.6 and 6.7), as well as 

market responses to changes in timber and agricultural production (U.S. EPA, 2014b, sections 

6.3 and 6.5). We additionally recognize the potential for O3 to impact other biomass-related 

responses, such as the supply of non-timber forest products and other ecosystem responses for 

which we have primarily qualitative characterizations of impacts (U.S. EPA, 2014b, chapter 5). 

We turn first to the WREA estimates for a range of effects related to biomass loss, which 

are based on application of the robust E-R functions for seedlings of 11 tree species described in 

the ISA (U.S. EPA, 2013, section 9.6.2) and the WREA (U.S. EPA, 2014b, Appendix 6F)8

. 

First, we note (as considered above) the range of responses for the individual species for which 

robust E-R functions have been developed. These eleven species vary appreciably in sensitivity 

of growth reduction (in terms of relative biomass loss, or RBL) in response to O3 exposure. 

 

8

 There is an E-R function available for a 12th tree species (cottonwood), but this E-R function is considered 

less robust because it is based on the results of a single gradient study (Gregg et al., 2003). That combined with its 

apparent extreme response to O3 prompted CASAC to advise the Administrator to not place too much emphasis on 

cottonwood in the review of the secondary standard (Frey, 2014, p. 10). As a result, while we do include 

cottonwood in some of the analyses, we have decided it would not be appropriate to put less weight on the 

cottonwood biomass loss estimates when considering what levels of W126 should be considered protective of 

median species biomass loss. 
6-19 

Based on the 11 individual tree species with robust seedling E-R functions, six of the 11 species 

show 2% seedling biomass loss at a W126 index value below 8 ppm-hrs and in the other five 

species at a W126 index value above 18 ppm-hrs. Within the group of six more sensitive 

species, the most sensitive is black cherry (see Figure 5-1B). 

 In Appendix 6F, the WREA presents individual and median response across the studied 

tree and crop species (U.S. EPA 2014, Appendix 6F). This appendix includes an analysis of the 

median of the composite exposure-response (E-R) functions for tree seedlings and crops. 

Specifically, Tables 6F-1 and 6F-2 provide estimates of the relative loss for trees and crops 

respectively at various W126 index values using the composite E-R functions for each species. 

The median of the composite functions is calculated for all 12 tree species as well as for the 11 

tree species excluding cottonwood. The median of the composite functions for all 12 tree species 

and all 10 crop species is consistent with the green line shown in Figures 6-5 and 6-6 (U.S. EPA 

2014, section 6.2.1.2. Tables 6F-3 and 6F-4 provide estimates of the number of species for trees 

and crops respectively that would be below various benchmarks (e.g., 2% biomass loss for trees) 

at various W126 index values. Based on the median composite E-R function developed for the 

11 tree species depicted in WREA Table 6F-1, median tree species biomass loss ranges from less 

than 1.5% to 5.3% over the W126 index value range of 7 to 17 ppm-hrs (U.S. EPA, 2014b, 

Appendix 6F). 

We additionally consider the WREA estimates of overall ecosystem-level effects from 

biomass loss considering the studied species together (U.S. EPA 2014, section 6.8). The WREA 

analysis used the species-specific biomass loss E-R functions, information on prevalence of the 

studied species across the U.S., and a weighting approach based on proportion of the basal area 

within each grid cell that each species contributes. The WREA analyses use information from 

the individual and median E-R functions for tree seedlings to provide information on the 

geographical extent of the effects of O3 exposure on growth reduction for different air quality 

scenarios. It provides information on the location and number of species affected, as well as 

information about the estimated effects in Class I areas. We note that some of these analyses 

continue to include cottonwood and where this is the case, it is so noted. In the WREA analyses, 

the largest reductions in O3 concentrations occur when air quality is adjusted from recent 

conditions to meeting the current standard. Smaller changes in O3 concentrations occur when air 

quality is adjusted for the W126 air quality scenarios for 15, 11, and 7 ppm-hrs (average across 

three years), relative to meeting the current standard. 

A weighted RBL value for each grid cell is generated by weighting the RBL value for 

each studied tree species found within that grid cell by the proportion of basal area it contributes 

to the total basal area of all (unstudied and studied) tree species within the grid cell, and then 
6-20 

summing those individual weighted RBLs. Table 6-2 below describes the percent of assessed 

geographic area with RBL exceeding 2% for 11 species based on the average W126 index values 

estimated for five air quality scenarios. Under recent conditions, 7.6 % of the total geographic 

area has a wRBL above 2% while just meeting the current standard across the contiguous U.S., 

the WREA estimates 0.2% of the total geographic area to have a weighted relative biomass loss 

above 2% for the 7 ppm-hrs scenario (Table 6-2 below; U.S. EPA 2014, Table 6-25). In the 

W126 air quality scenarios for 15, 11, and 7 ppm-hrs (average across three years), the percent of 

total area having weighted relative biomass loss greater than two percent was 0.2 percent, 0.1 

percent and <0.1 percent, respectively (Table 6-2 below; U.S. EPA 2014, Table 6-25). In 

considering these estimates, however, we note that the values for percentages of basal area 

include many grid cells in which none of the 11 studied species are found, and thus these values 

are likely to be low. In addition, the ecosystem level impacts from O3-induced effects on 

biomass loss in each grid cell would also depend on the interaction between the studied species 

with known O3-sensitivities and the other species that are also contributing to the total basal area 

which have unstudied O3-sensitivities. Given these and other potential uncertainties and 

limitations associated with this analysis (U.S. EPA, 2014b, section 6.8), which were also 

commented on by CASAC (Frey, 2014a, p. A-40), we thus conclude that this analysis does little 

to inform the nature and degree of risk likely to be experienced by O3-sensitive species growing 

in mixed-species forests, which are wide-spread in the eastern U.S. These values may be more 

appropriate for western forests which more often are composed of a single species (i.e., 

ponderosa pine, aspen forests). 

 

Table 6-2. Percent of assessed geographic area exceeding 2% weighted relative biomass 

loss in WREA air quality scenarios. 

Percent of total 

area with 

wRBL>2%

 ----------------Air Quality Scenarios --------------------------- 

Recent Conditions 

(2006-2008) 

Conditions just 

meeting the 

current standardA

W126 index scenariosB

15 ppm-hrs 11 ppm-hrs 7 ppm-hrs 

 Using all 12 

Species 10.8% 0.8 % 0.7 % 0.5 % 0.2 % 

Using 11 

species 

(excluding 

cottonwood) 

7.6% 0.2% 0.2% 0.1% <0.1%

A This analysis uses air quality values that are estimated per model grid cell using the W126 index value assigned to the grid 

cell based on application of the VNA method to the monitor-location W126 index values that are the average at that location 

across the 3 years of W126 index values for the adjusted dataset that just meets the current standard (4th highest daily 

maximum 8-hour concentration, averaged over 3 consecutive years of 75 ppb).
6-21 

B The national distribution of W126 index values within model grid-cells for each scenario reflects model-based adjustment of 

2006-2008 O3 concentrations at monitoring sites such that the average W126 index at the controlling location in each of the 

modeling regions just meets the scenario target index value, followed by application of the VNA interpolation methodology 

(see U.S. EPA, 2014b, section 4.3.4.1 and Appendix 4A).

To further inform this issue, the WREA characterized the number of counties where the 

median RBLs were above 2% (U.S. EPA, 2014b, Table 6-7), as shown in Table 5-5. Given 

CASAC’s advice to put less emphasis on cottonwood, we focus on the rows of this table that 

excluded cottonwood. Under recent conditions, 52% of the counties have median RBLs above 

2%. When air quality is adjusted to the current standard, that proportion drops to 8% and further 

decreases to 6% for air quality adjusted to just meet a 3-year average W126 level of 7 ppm-hrs. 

With respect to median RBL values, of the 239 counties (8% of counties) estimated to have a 

median RBL above 2% when meeting the current standard, 203 of those counties have a RBL 

above 2% because of the presence of black cherry. Thus, as also discussed above in Section 6.2, 

given the magnitude of estimated RBL for black cherry over the entire range assessed, it is not 

clear to what extent the information for black cherry informs the selection of an appropriate 

level. 

In addition, the WREA also characterized the number of counties where one or more 

individual studied tree species showed a 2% biomass loss (U.S. EPA, 2014b, Table 6-7), as also 

shown in Table 5-5. This is consistent with CASAC advice that “rather than focusing solely on 

the median relative biomass loss (RBL), the number of counties containing sensitive tree species 

that are expected to have growth loss of greater than 2% should be quantified” (Frey, 2014a, p. 

11). The maximum number of species that exceed 2% RBL in any one county is five species, 

which only occurs under recent O3 conditions. After meeting the current standard, the maximum 

number of species in any one county is four. This information shows that a number of counties 

have more than one O3-sensitive species growing in it, potentially together in the same forest 

stands, whose RBLs are above 2%. Given CASAC’s advice to put less emphasis on cottonwood, 

we focus on the rows of this table that excluded cottonwood. Under recent air quality conditions, 

the proportion of counties with 1 or more species with an RBL greater than 2% is 78% (2,418 

counties). As air quality is adjusted to just meet the current standard and the alternative W126 

index value of 7 ppm-hrs, this number drops to 62% and 58%, respectively. We note that of the 

1929 counties estimated to have 1 or more species with an RBL greater than 2% when meeting 

the current standard9

, 1805 of those counties are estimated to have black cherry as the only 

species estimated to experience this level of biomass loss. Thus, as also discussed above in 

Section 6.2, given the magnitude of estimated RBL for black cherry over the entire range 

 

9

 Excluding cottonwood, as discussed above. 
6-22 

assessed, it is not clear to what extent the information for black cherry informs the selection of an 

appropriate level. We next consider the wRBL estimates from the WREA analysis of 145 (of the 

155) federally designated Class I areas for which there was sufficient information regarding O3-

sensitive species (U.S. EPA, 2014b, section 6.8.1, Table 6-26, Appendix 6E). These 145 parks 

had at least one O3-sensitive tree species for which an E-R function for RBL was available. 

Using the E-R functions for the species found within each park, the WREA calculated an average 

wRBL value for each park for the 3-year average W126 index values estimated in those locations 

for the current standard and three W126 air quality scenarios. Under conditions adjusted to just 

meet the current standard, the average wRBL in 2 of the 145 parks is estimated to be above 2%, 

as presented in Table 6-3 below. These two parks are Badlands National Park, driven by 

sensitivity of cottonwood, and Wind Cave National Park, driven by sensitivity of ponderosa pine. 

We compare this estimate to those for the W126 scenarios. For the W126 scenarios of 15 and 11 

ppm-hrs, the estimated weighted RBL is greater than 2% in these same two of the 145 parks, 

while it is greater than 2% in only 1 park (Wind Cave) for the 7 ppm-hrs scenario. 

Table 6-3. Number of Class I areas (of 145 assessed) with weighted relative biomass loss 

greater than 2%. 

 ----------------Air Quality Scenarios --------------------------- 

Conditions just meeting 

the current standardA

3-Year Average W126 index scenariosB

15 ppm-hrs 11 ppm-hrs 7 ppm-hrs 

Number of Class I areas with 

wRBL>2% 2 2 2 1 

A The wRBL is estimated per model grid cell (in which there are any of the 12 studied species) from W126 index value assigned 

to the grid cell based on application of the VNA method to the monitor-location W126 index values that are the average at that 

location across the 3 years of W126 index values for the adjusted dataset that just meets the current standard (4th highest daily 

maximum 8-hour concentration, averaged over 3 consecutive years of 75 ppb). 

B The national distribution of W126 index values within model grid-cells for each scenario reflects model-based adjustment of 

2006-2008 O3 concentrations at monitoring sites such that the average W126 index at the controlling location in each of the 

modeling regions just meets the scenario target index value, followed by application of the VNA interpolation methodology (see 

U.S. EPA, 2014b, section 4.3.4.1 and Appendix 4A).

The WREA estimates of crop yield loss for the modeled air quality scenarios are 

summarized in Table 6-4 below (details are provided in U.S. EPA 2014, section 6.5.1 and 

Appendix 6B). For the recent air quality conditions scenario, the means for all crops were less 

than 5% loss across all states. Crop yield loss estimates for all states were also less than 5% in 

the air quality scenario representing conditions just meeting the current standard (U.S. EPAb, 

2014, section 6.5.1 and Appendix 6B). 
6-23 

Table 6-4. Estimated mean yield loss (and range across states) due to O3 exposure for two 

important crops. 

Crop 

------------------------------ Air Quality Scenarios ------------------------------------- 

Recent 

Conditions 

(2006-2008) 

Conditions just 

meeting the current 

standardA

Average W126 index scenariosB

15 ppm-hrs 11 ppm-hrs 7 ppm-hrs 

Corn 

<5% C

(0.01-0.88) 

<5% 

(0.0-0.01) 

<5% 

(0.0-0.01) 

<5% 

(0.0 – 0.0) 

<5% 

(0.0 – 0.0) 

Soybean 

<5% 

(0.69-8.30) 

<5% 

(0.01 – 1.39) 

<5% 

(0.01 – 1.13) 

<5% 

(0.01 – 0.75) 

<5% 

(0.01 – 0.59) 

A The crop yield loss is estimated per grid cell (and per FASOMGHG region) from W126 index value assigned to the cell based 

on application of the VNA method to the monitor-location W126 index values that are the average at that location across the 3 

years of W126 index values for the adjusted dataset that just meets the current standard (4th highest daily maximum 8-hour 

concentration, averaged over 3 consecutive years of 75 ppb). 

B The national distribution of W126 index values within grid cells for each scenario reflects model-based adjustment of 2006-

2008 O3 concentrations at monitoring sites such that the average W126 index at the controlling location in each of the 

modeling regions just meets the scenario target index value, followed by application of the VNA interpolation methodology (see 

U.S. EPA 2014 section 4.3.4.1 and Appendix 4A). 

C Mean yield loss is the mean across modeling units. The range presented in parentheses below the mean represents the 

minimum and maximum estimates across modeling units (U.S. EPA 2014, Appendix 6B).

The WREA also analyzes market responses to changes in timber and agricultural 

production (U.S. EPA, 2014b, sections 6.3 and 6.5). As explained above, however, comparisons 

of the WREA’s air quality scenarios for the national-scale estimates of timber production and 

consumer and producer surpluses are not straightforward to interpret due to market dynamics. 

Estimates for the recent conditions and current standard scenarios are compared to the three 

W126 scenarios. In general, substantially greater economic surpluses (approximately $51 

billion) are estimated from the comparison of the recent conditions (2006-2008) scenario to the 

current standard scenario. The vast majority of these economic surpluses are estimated for 

agricultural production. Differences of the average W126 scenarios from the current standard 

scenario are much smaller (U.S. EPA 2014, Appendix 6B). 

Because increases in timber production represent increased tree growth and concurrent 

carbon sequestration, we also consider WREA estimates of the potential increase in carbon 

storage that potentially could occur for different air quality scenarios (U.S. EPA 2014, section 

6.6.1). Comparisons of the W126 scenarios to the current standard scenario with regard to 

carbon sequestration estimates do not indicate an appreciable difference for the W126 scenario 

of 15 ppm-hrs beyond that achieved by just meeting the current standard. The majority of the 

enhanced carbon sequestration potential resulting from increases in forest biomass is predicted to 

occur for the W126 scenarios of 11 and 7 ppm-hrs. Over 30 years, the current standard scenario 
6-24 

projection is 89,184 million metric tons of CO2 equivalents (MMtCO2e).10 The WREA estimates 

additional sequestration potential of 13, 593 and 1,600 MMtCO2e, for the W126 scenarios of 15, 

11 and 7 ppm-hrs, respectively, as compared to the current standard (U.S. EPA 2014, Table 6-

18). We also take note of the relatively smaller estimates for carbon sequestration associated 

with improved crop yields (over 30 years) in the agricultural sector, which indicate little 

difference among the different W126 scenarios, beyond that achieved by just meeting the current 

standard. 

We additionally consider the WREA estimates for five urban areas of how reduced 

growth of O3-sensitive trees in urban forests may affect the ecosystem services of air pollutant 

removal and carbon sequestration (U.S. EPA, 2014b, sections 6.6.2 and 6.7 and Appendix 6D). 

With regard to air pollutant removal, the WREA estimated metric tons of carbon monoxide, 

nitrogen dioxide, ozone and sulfur dioxide removed under the W126 scenarios. In considering 

these estimates we note the general assumptions made to estimate order of magnitude effects of 

O3 removal by trees on O3 concentrations in the five urban areas and the associated uncertainties 

(U.S. EPA 2014, sections 6.7 and 6.9 and Appendix 6D). Estimates for all five case study areas 

indicate increased pollutant removal from the recent conditions scenario to just meeting the 

current standard scenario, with much smaller differences between the current standard and the 

three W126 scenarios (Table 6-5 below). The largest difference in carbon sequestration is 

between the existing conditions scenario and the current standard scenario (Table 6-5). In 

addition to the small differences in W126 index values among the three W126 air quality 

scenarios relative to the current standard for these five areas, only 2 or 3 tree species were able to 

be assessed in each city. Therefore, these results may underestimate the overall impacts in these 

areas and nationally, although other areas of uncertainty (recognized below) may tend to 

contribute to the opposite potential (U.S. EPA 2014, Table 6-27).

 

10 1 MMtCO2e is equivalent to 208,000 passenger vehicles or the electricity to run 138,000 homes for 1 

year as calculated by the EPA Greenhouse Gas Equivalencies Calculator (updated September 2013 and available at 

http://www.epa.gov/cleanenergy/energy-resources/calculator.html). 
6-25 

Table 6-5. Estimated effect of O3-sensitive tree growth-related impacts on the ecosystem 

services of air pollutant removal and carbon sequestration in five urban case 

study areas. 

Case Study Area 

------------------------------------------ Air Quality Scenarios ----------------------------- 

Recent 

Conditions 

(2006-2008) 

Conditions just 

meeting the 

current 

standardA

Average W126 index scenariosB

15 ppm-hrs 11 ppm-hrs 7 ppm-hrs 

Air Pollutant Removal (metric tons, CO, NO2, O3, SO2)

Atlanta 33,000 35,800 35,800 36,000 36,300 

Baltimore 8,500 9,200 9,200 9,200 9,200 

Chicago 355,000 359,000 359,000 361,000 365,000 

Syracuse 1,500 1,700 1,700 1,700 1,700 

Tennessee urban 474,000 511,000 511,000 516,000 522,000 

Carbon Storage (million metric tons of CO2 equivalents, cumulative over 25 years)

Atlanta 1.2 1.32 1.32 1.32 1.34 

Baltimore 0.5 0.57 0.57 0.57 0.57 

Chicago 16.9 17.05 17.05 17.10 17.21 

Syracuse 0.14 0.17 0.17 0.17 0.17 

Tennessee urban 18.0 19.67 19.67 19.89 20.16 

A Results are derived from estimates per model grid cell (in which there are any of the 12 studied species) from W126 index 

value assigned to the grid cell based on application of the VNA method to the monitor-location W126 index values that are 

the average at that location across the 3 years of W126 index values for the adjusted dataset that just meets the current 

standard (4th highest daily maximum 8-hour concentration, averaged over 3 consecutive years of 75 ppb). 

B The national distribution of W126 index values within model grid-cells for each scenario reflects model-based adjustment of 

2006-2008 O3 concentrations at monitoring sites such that the average W126 index at the controlling location in each of the 

modeling regions just meets the scenario target index value, followed by application of the VNA interpolation methodology 

(see U.S. EPA, 2014b, section 4.3.4.1 and Appendix 4A).

With regard to foliar injury, we take note of the WREA analyses of the nationwide 

dataset (2006- 2010) for USFS/FHM biosites described in section 5.4.2 above, including the 

observation that the proportion of biosites with injury varies with soil moisture conditions and O3

W126 index values (U.S. EPA 2014, Chapter 7, Figure 7-10). The evidence of O3-attributable 

visible foliar injury incidence occurring in USFS/FHM biosites shows that the proportion of 

biosites showing foliar injury incidence increases steeply with W126 index values up to 

approximately 10 ppm-hrs. At W126 index levels greater than approximately 10 ppm-hrs, the 

proportion of sites showing foliar injury incidence is relatively constant.

In reflecting across the range of W126 index values evaluated in various WREA 

analyses, we first note the substantial reductions in biomass-related risks estimated for air quality 
6-26 

adjusted to just meet the current standard scenario. Additional incremental risk reductions are 

estimated across the W126 scenarios, although these risk reductions are substantially smaller. 

In considering the WREA estimates here, we take note of uncertainties in the adjusted 

estimates. Adjustments were made to recent air quality to reflect just meeting the current 

standard and three W126 levels. These adjustments were based on air quality modeling 

simulations reflecting across-the-board reductions in NOx emissions required to bring the highest 

monitor down to the target level in different regions of the country. In some areas, meeting a target 

level at the highest monitor in the region had the effect of substantially reducing concentrations 

below the targeted level in other parts of the region. This adjustment approach is not meant to 

represent an actual control strategy but to provide an approximation of the spatial variability of O3

across an area when just meeting the current standard and three W126 levels. 

We also note potential uncertainties in the extent to which the results for each modeled 

air quality scenario represent cumulative seasonal O3 exposures that would be expected to occur 

across the three years represented in each scenario. In general, each scenario is represented by a 

dataset of 3-year average W126 index values across the national modeling area. Thus, the results 

estimated for the various analyses that use these scenarios do not reflect any year-to-year 

variability that would be expected in single year results. Rather, they reflect average estimates 

for the three year period modeled. Analyses in the WREA describe the potential for the WREA 

estimates to underestimate cumulative biomass-related effects in perennial species (as noted in 

sections 6.2 and 5.2.2 above and described in detail in U.S. EPA, 2014b, chapter 6, 6.2.1.4). This 

potential for underestimation is recognized in the context of the uncertainties associated with 

other aspects of the different analyses in section 6.9 of the WREA (e.g., U.S. EPA, 2014b, Table 

6-27). We additionally note that the WREA compounding analyses do not take into account 

other variables that can affect the magnitude of these effects in the field. In considering this 

information discussed above in the context of identifying levels appropriate to consider for a 

W126-based standard, we take note of additional associated uncertainties as discussed under the 

following question. 

 What are important uncertainties and limitations in the evidence and 

exposure/risk analyses? 

In considering the evidence and exposure/risk information summarized above and the 

weight to place on this information, we are mindful of the uncertainties and limitations 

associated with several key aspects of this information. We first consider the uncertainties 

associated with the evidence underlying the tree seedling and crop E-R functions, given the 

importance of these functions for many of the ecosystem service analyses described in the 
6-27 

WREA. Several key uncertainties associated with this information are listed below and 

described in more detail in the WREA (U.S. EPA, 2014b). 

 Uncertainty regarding the extent to which the subset of studied tree and crop species 

encompass the total number of O3 sensitive species in the nation and to what extent it 

is representative of U.S. vegetation as a whole, given that information is available for 

only a small fraction of the number of total species of trees and crops grown in the 

U.S. (U.S. EPA, 2013, section 9.6, U.S. EPA, 2014b, Table 6-27). 

 Uncertainties regarding intra-species variability due to the different numbers of studies 

that exist for different species so that the weight of evidence is not the same for each 

species. Those species with more than one study show variability in response and E-R 

functions. The potential variability in less well-studied species is, however, unknown 

(U.S. EPA, 2013, pp. 9-123/125, U.S. EPA, 2014b, section 6.2.1.2, and Table 6-27). 

 Uncertainty regarding the extent to which tree seedling E-R functions can be used to 

represent mature trees since seedling sensitivity has been shown in some cases to not 

reflect mature tree O3 sensitivity in the same species (U.S. EPA, 2013, section 9.6, 

U.S. EPA, 2014b, section 6.2.1.1 and Tables 6-5 and 6-27). 

 Uncertainty in the relationship of O3 effects on tree seedlings (e.g., relative biomass 

loss) in one or a few growing seasons to effects that might be expected to accrue over 

the life of the trees extending into adulthood (U.S. EPA, 2013, pp. 9-52/53, U.S. 

EPA, 2014b, section 6.2.1.4 and Table 6-27). 

 Uncertainties associated with estimating the national scale ecosystem-level impacts 

using weighted relative biomass loss (U.S. EPA, 2014b, section 6.8, and Table 6-27). 

 Uncertainties associated with potential biomass loss in federally designated Class I 

areas (U.S. EPA, 2014b, section 6.8. and Table 6-27). 

Turning to consideration of the air quality conditions estimated for the various air quality 

scenarios, we take note of the following uncertainties associated particularly with estimates of O3

exposures in rural areas nationally. These are described more completely in chapter 4 of the 

WREA (see for example, U.S. EPA, 2014b, section 4.4) and summarized in chapter 8 of the 

REA (U.S. EPA, 2014b, section 8.5). 

 Uncertainties in O3 exposures due to a lack of rural monitors, especially in the western 

U.S. and at high elevation sites. 

 Uncertainties associated with the method (VNA) used to interpolate monitor values to 

estimate W126 index values in locations without monitors. 

 Uncertainties in adjusted estimates of O3 concentrations associated with meeting the 

current standard and potential alternative W126-based standards. 

Numerous ecosystem services assessments were described in the WREA. These 

assessments relied heavily on models, which also relied on the inputs of the tree seedling and 

crop E-R functions and adjusted air quality estimates. Thus, including the uncertainties from the 
6-28 

first two categories discussed above, additional uncertainties associated with the ecosystem 

services models include the following. 

 Uncertainties associated with use of the i-Tree model to estimate pollution removal 

and carbon storage in five urban area case studies, including uncertainties in the base 

inventory of city trees, the functions used for air pollution removal and carbon storage 

(U.S. EPA, 2014b, sections 6.6.2, 6.7, and Table 6-27). 

 Uncertainties associated with use of the FASOMGHG model for national timber and 

crop production, including use of median E-R functions for crops in FASOM and 

crop proxy and forest type assumptions to fill in where there was insufficient data 

(U.S. EPA, 2014b, sections 6.3, 6.5, 6.6.1, and Table 6-27). 

 Uncertainties associated with use of the FASOMGHG model to estimate national scale 

carbon sequestration, including those associated with the functions for carbon 

sequestration (U.S. EPA, 2014b, sections 6.2.1.1, 6.6.1, and Table 6-27). 

In addition, the WREA estimates the incidence and of O3-induced visible foliar injury, 

both at the national and national park scales. Numerous uncertainties are associated with these 

assessments and include the following. 

 Uncertainties associated with our understanding of the number and sensitivity of O3

sensitive species (U.S. EPA, 2014b, sections 7.2.1, 7.5 and Table 7-22). 

 Uncertainties associated with spatial assignment of foliar injury biosite data to 12x12 

km grids (U.S. EPA, 2014b, sections 7.2.1, 7.5 and Table 7-22). 

 Uncertainties associated with availability of biosite sampling data in some locations in 

the western U.S. (U.S. EPA, 2014b, sections 7.2.1, 7.5 and Table 7-22). 

 Uncertainties associated with soil moisture threshold for foliar injury (U.S. EPA, 

2014b, sections 7.2.2, 7.2.3, 7.5 and Table 7-22). 

 Uncertainties associated with spatial resolution of soil moisture data, time period for 

soil moisture data, drought categories and the combination of soil moisture and 

biosite data (U.S. EPA, 2014b, sections 7.3.3.2, 7.5 and Table 7-22). 

 Uncertainties associated with O3 exposure data of vegetation and recreational areas 

within parks (U.S. EPA, 2014b, sections 7.4, 7.5 and Table 7-22). 

 Uncertainties associated with surveys of recreational activities (U.S. EPA, 2014b, 

sections 7.1.1.2, 7.5 and Table 7-22). 

Additionally, there is uncertainty associated with the extent to which the endpoints and 

associated risk estimates considered above represent effects reasonably judged adverse in the 

context of public welfare. Despite these uncertainties, the overall body of scientific evidence 

underlying the ecological effects and associated ecosystem services evaluated in the WREA is 

strong, and the methods used to quantify associated risks are scientifically sound (Frey, 2014b). 
6-29 

All of these uncertainties are important to considerations below in the context of target levels of 

protection with regard to weight to be placed on various lines of evidence and assessment results. 

 Are there other aspects of the form that affect consideration of the welfare 

protection provided by the level of the cumulative seasonal standard? 

Although cumulative, seasonal exposure indices of interest for vegetation effects are 

often expressed in terms of a single season, we recognize that it can also be appropriate to 

consider a form that is evaluated over a multiple-year period, such as three years (U.S. EPA, 

2007; 72 FR 37901; 75 FR 3021). The current form of the secondary standard is a 3-year 

average, and we recognize that the protection provided by the secondary standard derives from 

the combination of all elements of the standard (indicator, form, averaging time(s), and level). 

Thus, we find it appropriate to evaluate the protection that might be afforded by a form limited to 

a single year or one that is based on evaluation of exposures across multiple years. Although 

cumulative, seasonal exposure indices of interest for vegetation effects are often expressed in 

terms of a single season, we recognize that it can also be appropriate to consider a form that is 

evaluated over a multiple-year period, such as three years (U.S. EPA, 2007; 72 FR 37901; 75 FR 

3021). Accordingly, this discussion explores the information relevant to consider in conjunction 

with the above identification of the W126 index form, 12-hour daylight averaging time and 

maximum consecutive 3-month seasonal exposure period, and the subsequent discussion on level 

below, when considering support in the current information for single and/or multiple-year 

options. 

We additionally take note of advice from CASAC on this topic in the current and prior 

reviews. Specifically, in this review, CASAC stated that it “does not recommend the use of a 

three-year averaging period for the secondary standard. We favor a single-year period for 

determining the highest three-month summation which will provide more protection for annual 

crops and for the anticipated cumulative effects on perennial species. The scientific analyses 

considered in this review, and the evidence upon which they are based, are from single-year 

results. If, as a policy matter, the Administrator prefers to base the secondary standard on a threeyear averaging period for the purpose of program stability, then the level of the standard should 

be revised downward such that the level for the highest three-month summation in any given 

year of the three-year period would not exceed the scientifically recommended range of 7 ppmhrs to 15 ppm-hrs” (Frey, 2014a, p. iii). 

In considering an annual form of a standard, we particularly take note of O3-induced 

vegetation effects that can occur as a result of a single year’s exposure. These include visible 

foliar injury symptoms, growth reduction in annual and perennial species, and yield loss in 

annual crops. The following discussion considers these effects, in the context of their potential 

public welfare significance, and in regard to the extent to which a W126-based standard with an 
6-30 

annual form or one based on evaluation across multiple years may be able to provide appropriate 

protection 

In the case of foliar injury, the ISA notes that the full body of evidence indicates that 

there is wide variability in this endpoint, such that although evidence shows visible foliar injury 

can occur under very low cumulative O3 concentrations, “...the degree and extent of visible 

foliar injury development varies from year to year and site to site... even among co-members of 

a population exposed to similar O3 levels, due to the influence of co-occurring environmental 

and genetic factors” (U.S. EPA 2013, section 9.4.2, p. 9-38). In addition, the WREA assessment 

of foliar injury showed the difficulty and complexity associated with identifying W126 index 

values that would consistently provide appropriate protection on an annual basis for this 

endpoint. We thus conclude that there is limited information to discern between the level of 

protection provided by an annual form or a 3-year average form of a W126 standard for this 

endpoint, and that a multiple year form could be considered to provide a more consistent target 

level of protection for this endpoint, given likely fluctuations in annual O3 and soil moisture 

conditions. 

In the case of annual commodity crops, the overall welfare effect of annual changes in 

yields due to O3 exposures is not straightforward. As noted above, determining at what point O3-

induced crop yield loss becomes adverse to the public welfare is still unclear, given that it is 

heavily managed with additional inputs that have their own associated markets and that benefits 

can be unevenly distributed between producers and consumers. We thus conclude that there is 

limited information to discern between the level of protection provided by an annual form or a 3-

year average form of a W126 standard for this endpoint. As with foliar injury, we thus conclude 

that it is appropriate to consider a level of protection for annual commodity crops that would be 

achieved, on average, using a multiple year form, to provide a more consistent target, given 

likely fluctuations in environmental and economic conditions. 

In contrast to impacts on annual species that accrue in the single growing season in which 

the O3 exposures occur, annual effects in perennial species can be “carried over” into the 

subsequent year where they affect growth and reproduction (U.S. EPA, 2013, pp. 9-43 to 9-44 

and p. 9-86). In addition, when these effects occur over multiple years due to elevated O3

exposures across several years, they accumulate and potentially compound, increasing the 

potential for effects at the ecosystem level and associated ecosystem services that may be of 

significance to the public welfare. 

Effects from elevated O3 years on perennial plants, when they occur over several years, 

can be propagated up to higher spatial scales where they can contribute to effects on ecosystem 

services, e.g., alteration of below-ground biogeochemical cycles, and alteration of both aboveand below- ground terrestrial community composition and terrestrial ecosystem water cycling 
6-31 

(U.S. EPA, 2013, Table 9-19). Ozone has also been shown to affect plant reproduction in 

numerous ways (U.S. EPA, 2007, 7.3.3.3; U.S. EPA, 2013, 9.4.3.1). These effects, when they 

occur at sufficient magnitude for a single species, may result in impaired recruitment and loss of 

the species from the stand or community. This has the potential to change the community 

composition and biodiversity. If these effects occur in multiple plant species and/or over 

multiple years, they can result in a reduction in the productivity and carbon sequestration of 

terrestrial ecosystems. Such ecosystem-related effects and others discussed in the ISA may be 

considered to reflect impacts of critical O3 exposures over the longer term. We additionally note 

that as compared to intermittent (or single year) critical O3 exposures, multiple years of such

exposures might be expected to result in larger impacts on forested areas, e.g., increased 

susceptibility to other stressors such as insect pests, disease, co-occurring pollutants and harsh 

weather, due to the potential for compounding or carry-over effects on tree growth. 

Given the above, we find it reasonable to conclude that the public welfare significance of 

the effects that can accumulate as a result of multiple-year O3 exposures have the potential to be 

greater and more certain than those that are realized in an individual year. Thus, to the extent 

that the focus for public welfare protection is on long-term effects that occur in sensitive tree 

species in natural forested ecosystems, including in federally protected areas such as Class I 

areas or on lands set aside by states, tribes and public interest groups to provide similar benefits 

to the public welfare, a cumulative seasonal standard that evaluates exposures across multiple 

years (in combination with an appropriate level) might be a more appropriate match to provide 

the requisite protection for those O3-related effects on vegetation that when accumulated across 

years, are potentially significant and adverse to the public welfare. 

Additionally, we address the potential for cumulative impacts on biomass loss over a 3-

year period versus a 1-year period. First it is important to note that the WREA analyses that 

characterize plant biomass and associated ecosystem services effects, discussed above in this 

section, are based on a 3-year average. The WREA analysis examined the potential for biomass 

loss estimates based on a 3-year average W126 index value to underestimate the cumulative 

impact on growth based on the biomass loss that would be predicted in each of the 3 years, based 

on the yearly W126 index values. The results show that the use of the three-year average W126 

index value may underestimate RBL values slightly. However, it should be noted that the 

approach does not account for moisture levels or other environmental factors that could affect 

biomass loss (U.S. EPA, 2014b, section 6.2.1.4 and Figure 6-14). In considering these results, 

we note that in these regions and in all three years, the three-year average W126 index value is 

sometimes above and sometimes below the individual year W126 index value. 

In addition to the vegetation effects considerations described above, there are other 

policy-relevant factors that can be useful to consider. For example, under a standard with a 
6-32 

single year form, a monitor may be judged to meet the standard based on a single year of data, 

while under a standard with a form requiring evaluation over a multi-year period, a monitor is 

not judged to have met the standard until a complete multi-year record is available. For a W126-

based potential standard, the multi-year form identified for consideration in the last review was 

the average cumulative seasonal metric over three consecutive years (75 FR 3027). Such a 

multi-year form remains appropriate to consider to provide stability to an alternative secondary 

standard, just as the multi-year form provides for the current standard (average over three years 

of annual fourth-highest daily maximum 8-hour average O3 concentrations).11 In considering the 

issue of stability in the context of such a form, we first note the inter-annual variability of 

seasonal W126 index, which is not unexpected given the logistic weighting function and also 

inter-annual variability in meteorological conditions that contribute to O3 formation (see 

Appendix 2C). The staff analysis in Appendix 2C describes the variability in annual W126 

index values in relation to variability in the 3-year average, which indicates that a standard based 

on an annual W126 index would be expected to have a lower degree of year-to-year stability 

relative to a standard based on a form that averages seasonal indices across three consecutive 

years. A more stable standard can be expected to contribute to greater public welfare protection 

by limiting year-to-year disruptions in ongoing control programs that would occur if an area was 

frequently shifting in and out of attainment due to extreme year-to-year variations in 

meteorological conditions. This greater stability in air quality management programs thus 

facilitates achievement of the protection intended by a standard. In light of this relationship, we 

conclude that a 3-year average form has the desirable feature of providing greater stability in air 

quality management programs and thus facilitating the achievement of the protection intended by 

a standard. Thus, we recognize the public welfare benefits of having a standard of a 3-year 

average form. 

CASAC has asked that the PA quantify the ratio of the 3-year average of the highest 

three-month summations in each year to the highest three-month summation in the highest year 

within that same 3-year average period. This information is provided in a technical 

memorandum titled “Relationship between W126 annual values and three-year averages” (EPAHQ-OAR-2005-0172) and in the analyses included in Appendix 2C. In the technical 

memorandum, an analysis summarized the relationship between annual W126 index values and 

the three-year averages of the annual values based on 2007-2009 air quality data. Based on the 

air quality data, 79 percent of counties meeting a three-year average W126 index value of 13 

ppm-hrs would also not have annual W126 index values above 15 ppm-hrs. In addition, in terms 

 

11 See ATA III, 283 F. 3d at 374-75 (recognizing programmatic stability as a legitimate consideration in the 

NAAQS standard-setting process). 
6-33 

of county-years (i.e., the number of counties times the number of years in the analysis), 93 

percent of the county-years meeting a three-year average W126 index value of 13-ppm-hrs 

would also meet an annual W126 index value of 15 ppm-hrs. 

In addition, Appendix 2C compared annual W126 index values to three-year average 

W126 index values for 2008-2010 air quality data. It concluded that the data analysis “shows that 

the inter-annual variability in the annual W126 index tends to decrease with decreasing W126 

levels. Thus, it is expected that reductions in NOx emissions will not only result in lower 3-year 

average W126 levels, but also result in less inter-annual variability associated with annual W126 

levels.” Appendix 2C also concludes that the inter-annual variability in the W126 index increases 

and decreases along with the three-year average. 

These analyses suggest that meeting a 3-year average W126 index value of 13 ppm-hrs 

would mean that for most years and monitoring sites the annual W126 index value would be 

below 15 ppm-hrs. In addition, the relationship between 3-year average W126 index values and 

annual W126 index values is dynamic and varies with the three-year average W126 index value 

and will continue to change in the future with changing pollution levels. 

Accordingly, in considering all elements for a revised standard, including level and form, 

we note that a standard with a form that averages across three years can also control for year-toyear variability and individual year concentrations. The appropriate level and form combination 

will depend on which effects endpoints are considered to warrant additional public welfare 

protection and what is considered to be the requisite range of target levels of protection. In 

articulating these objectives it may be appropriate to evaluate the nature of the O3 induced effects 

and their significance or importance to the public welfare, as well as the role that year-to-year 

exposure variability can play in public welfare impacts. 

 What considerations may be important to the Administrator’s judgments on the 

public welfare significance of O3 associated vegetation effects that may be 

expected under air quality conditions associated with different levels for a 

seasonal cumulative standard? 

Our consideration of this question is intended to provide a public welfare context for 

consideration of the evidence and exposure/risk information discussed above, which includes the 

nature and magnitude of observed and predicted effects at various levels of cumulative seasonal 

exposures. We also note the importance of considering information in an integrated manner, 

rather than focusing only on results from any one analysis. For example, we find it appropriate, 

in considering the evidence with regard to seedling growth reduction (or biomass loss), to 

consider the WREA estimates of affected area based on tree basal area together with estimates of 

individual species responses based simply on the evidence-based E-R functions, and in light of 

other potential impacts summarized above. In so doing in section 6.5 below, we take into 
6-34 

account considerations relevant to public welfare policy judgments required of the 

Administrator, such as those described here. 

As recognized in sections 1.3.2 and 5.1, the Clean Air Act specifies that secondary 

standards specify a level of air quality that is requisite to protect against known or anticipated 

adverse effects to public welfare. In the Administrator’s judgment as to the standards that would 

be requisite (i.e., neither more nor less stringent than necessary) to protect the public welfare 

under the Act, she may consider a number of factors including 1) what should be considered to 

constitute an adverse effect to the public welfare; 2) the nature and magnitude of the effects and 

the risks that remain after meeting the level of the current standard; and, 3) what is necessary to 

achieve the requisite (no more and no less) degree of public welfare protection. In the 2008 

decision by which the current standard was established, the Administrator considered these 

factors in judging the previously existing standard to not provide the requisite public welfare 

protection. At that time the Administrator found that the exposure- and risk-based analyses 

available in that review indicated that adverse effects to vegetation would be predicted to occur 

under air quality conditions associated with just meeting the then-current standard. The effects 

identified were “visible foliar injury and seedling and mature tree biomass loss in O3-sensitive 

vegetation” (73 FR 16496). In so noting, the Administrator indicated that he believed that “the 

degree to which such effects should be considered to be adverse depends on the intended use of 

the vegetation and its significance to public welfare” (73 FR 16496). With regard to 

consideration of intended use, the Administrator took note of the specific uses of public lands set 

aside by Congress and intended to provide benefits to the public welfare, “including lands that 

are to be protected so as to conserve the scenic value and the natural vegetation and wildlife 

within such areas, and to leave them unimpaired for the enjoyment for future generations” such 

as Class I areas (73 FR 16496). The Administrator also recognized areas set aside by states, 

tribes and public interest groups with the intent “to provide similar benefits to the public welfare, 

for residents on State and Tribal lands, as well as for visitors to those areas” (73 FR 16496).12 

In the Administrator’s judgments in the 2008 review, he did not identify specific criteria 

or benchmarks or a specific level of protection from adverse environmental effects to public 

welfare judged to be requisite under the Act.13 As noted above, the scientists at the 1996 

 

12 In considering areas that have not been afforded such special protection, ranging from vegetation used 

for residential or commercial ornamental purposes, such as land use categories that are heavily managed for 

commercial production of commodities such as agricultural crops, timer and ornamental vegetation, the 

Administrator indicated his expectation that protection of sensitive natural vegetation and ecosystems might be 

expected to also provide some degree of additional protection for heavily managed commercial vegetation (73 FR 

16496). 13 In remanding the 2008 decision on the secondary standard back to the EPA, the Court of Appeals for the 

D.C. Circuit determined that EPA did not specify what level of air quality was requisite to protect public welfare 
6-35 

workshop identified ranges of cumulative seasonal index values (e.g., in terms of SUM06 or 

W126) in the context of considering a degree of protection for vegetation effects defined in terms 

of relative yield loss in crops and relative biomass loss in tree seedlings. Considering this 

information in the context of a secondary standard entails policy judgments by the Administrator 

with regard to the degree that impacts exceeding these or other benchmarks and other effects 

should be judged adverse to the public welfare. In considering levels for a W126-based 

secondary standard that may be appropriate to consider, we recognize that the statute requires 

that a secondary standard be protective against only those known or anticipated O3 effects that 

are “adverse” to the public welfare, not all identifiable O3-induced effects. Thus, we recognize 

both the importance of scientific consensus statements that have been made regarding 

vegetation-related endpoints and O3 exposure levels that might protect against such key 

endpoints and the importance of placing such conclusions in the context of consideration of the 

public welfare more broadly. 

As discussed in section 5.1 and recognized by the EPA in prior reviews, staff recognizes 

the importance of a more expansive construct or paradigm that addresses what constitutes 

adverse effects of O3 to public welfare. In so doing, we also recognize several aspects or 

dimensions of vegetation effects for consideration within this paradigm. These include the 

likelihood, type, magnitude, and spatial scale of the effect, as well as the potential for recovery 

and any uncertainties relating to these conditions (77 FR 20231). As in the last review, we also 

continue to recognize that the public welfare significance of O3-induced effects on sensitive 

vegetation growing within the U.S. can vary, depending on the nature of the effect, the intended 

use of the sensitive plants or ecosystems, and the types of environments in which the sensitive 

vegetation and ecosystems are located. Any given O3-related effect on vegetation and 

ecosystems (e.g., biomass loss, foliar injury), therefore, may be judged to have a different degree 

of impact on the public welfare depending, for example, on whether that effect occurs in a Class 

I area, a city park, or in commercial cropland. In the 2008 review, the Administrator judged it 

appropriate that this variation in the significance of O3-related vegetation effects should be taken 

into consideration in judging the level of ambient O3 that is requisite to protect the public welfare 

from any known or anticipated adverse effects (73 FR 16496). For example, in considering 

visible foliar injury and seedling and mature tree biomass loss in O3-sensitive vegetation 

expected under alternative air quality scenarios, the Administrator noted that “the degree to 

which such effects should be considered to be adverse depends on the intended use of the 

vegetation and its significance to the public welfare” (73 FR 16496). Further, the rulemaking 

 

from adverse public welfare effects or explain why any such level would be requisite, as described in section 1.2.2 

above. Mississippi, 744 F.3d at 272-73. 
6-36 

notice stated that “[i]n considering what constitutes a vegetation effect that is adverse from a 

public welfare perspective, the Administrator believes it is appropriate to continue to rely on the 

definition of ‘adverse,’ ... that imbeds the concept of ‘intended use’ of the ecological receptors 

and resources that are affected, and applies that concept beyond the species level to the 

ecosystem level” (73 FR 16496). The notice went on to state that “[i]n so doing, the 

Administrator has taken note of a number of actions taken by Congress to establish public lands 

that are set aside for specific uses that are intended to provide benefits to the public welfare, 

including lands that are to be protected so as to conserve the scenic value and the natural 

vegetation and wildlife within such areas, and to leave them unimpaired for the enjoyment of 

future generations” (73 FR 16496). Such public lands that are protected areas of national interest 

include national parks and forests, wildlife refuges, and wilderness areas. 

We also consider effects on ecosystem services in considering adversity to public 

welfare. For example, the WREA has evaluated the economic value of ecosystem services 

affected by O3 and how those services might be expected to change under different air quality 

scenarios representing the current and potential alternative standards (U.S. EPA, 2014b, chapters 

6 and 7). 

Lastly, we recognize several important considerations in evaluating levels of protection 

and levels for a cumulative seasonal W126-based standard including: the extent of areas 

expected to be affected nationwide and the magnitude of those effects; the extent of effects in 

areas of national significance; the extent to which these impacts might be judged significant from 

a public welfare perspective and associated uncertainties in the information. Accordingly, we 

recognize that the range of alternative standard levels that may be appropriate to consider differs 

based on the weight placed on different aspects of the evidence and on different aspects of the 

quantitative exposure/risk information, and the associated uncertainties, as well as on public 

welfare policy decisions regarding the public welfare significance of the effects considered and 

the approaches for considering benchmarks for growth or biomass loss and other vegetation 

effects of O3. As described in chapter 1, our objective is to identify the range of policy options 

supported by the current evidence- and exposure/risk-based information and with consideration 

of the role of the Administrator’s public welfare judgments. In so doing, we recognize support 

for consideration of a broad range of W126 index values, which we discuss in section 6.5, with 

recognition of the different judgments that might provide support for different parts of such a 

range. 
6-37 

6.4 CONSIDERATION OF PROTECTIVENESS OF REVISED PRIMARY 

STANDARD 

In staff consideration of the primary standard in chapter 4, staff concludes it is 

appropriate to consider alternative primary standards of the same form and averaging time as the 

current primary standard and a lower standard level within the range of 60 to 70 ppb. Thus, 

although the discussion in this chapter, with regard to the secondary standard, indicates the 

appropriateness of considering an alternative secondary standard with a cumulative, seasonal 

form, we also recognize that, to the extent that the Administrator may find it effective to control 

air quality using the same form for both the primary and secondary standards, it may be practical 

to consider the extent to which a standard in the form of the primary standard might be expected 

to also reduce and provide protection from cumulative seasonal exposures of concern. For 

example, if a clear and robust relationship was found to exist between 8-hour daily peak O3

concentrations and cumulative, seasonal exposures, the averaging time and form of the current 

standard might be concluded to have the potential to be effective as a surrogate. In response to 

this, we ask the following question: 

 What does the available information indicate with regard to protection of welfare 

from cumulative O3 exposures that might be afforded by alternative secondary 

standards based on the form of the current standard (a 3-year average of 4th

highest 8-hour average concentrations)? 

Addressing this point, the ISA describes the results of a recent focus study that examined 

the diel14 variability in O3 concentrations in six rural areas between 2007 and 2009 (U.S. EPA, 

2013, pp. 3-131 to 3-133). The ISA reported that “[t]here was considerable variability in the diel 

patterns observed in the six rural focus areas” with the three mountainous eastern sites exhibiting 

a “generally flat profile with little hourly variability in the median concentration and the upper 

percentiles”, while the three western rural areas demonstrated a “clear diel pattern to the hourly 

O3 data with a peak in concentration in the afternoon similar to those seen in the urban areas”, 

which was especially obvious at the San Bernardino National Forest site, 90 km east of Los 

Angeles at an elevation of 1,384 meters (U.S. EPA, 2013, p. 3-132). Thus, while the western 

sites that are influenced by upwind urban plumes may have increased cumulative seasonal values 

coincident with increased daily 8-hour peak O3 concentrations, this analysis indicates that, in 

sites without such an urban influence (the eastern sites in this analysis), such a relationship does 

not occur (U.S. EPA, 2013, section 3.6.3.2). Thus, the lack of such a relationship indicates that 

in some locations, O3 air quality patterns can lead to elevated cumulative, seasonal O3 exposures 

without the occurrence of elevated daily maximum 8-hour average O3 concentrations (U.S. EPA, 

 

14 involving a 24-hr period 
6-38 

2013, section 3.6.3.2). Further, staff notes that the prevalence and geographic extent of such 

locations is unclear, since as in the last review, there continue to be relatively fewer monitors in 

the West, including in high elevation remote sites. In considering the findings of this analysis, 

we additionally recognize, however, that the cumulative seasonal values for the eastern rural 

sites, where cumulative seasonal O3 concentrations appear to be relatively less related to daily 

maximum 8-hour concentrations, are lower in general than those of the western, urbaninfluenced sites. 

In addition to the focus study described in the ISA (U.S. EPA, 2013, section 3.6.3.2), we 

considered analyses of air quality monitoring data and air quality modeling analyses. Chapter 2 

of this document characterizes recent monitoring data on O3 air quality in rural areas. While 

approximately 80% of the O3 monitoring network is urban focused, about 120 rural monitors are 

divided among CASTNET, NCore, and portable ozone monitors (POMs) sites (Chapter 2, pp. 2-

2 to 2-3, Figure 2.1). Specifically, as stated in chapter 2 “[a]lthough rural monitoring sites tend 

to be less directly affected by anthropogenic pollution sources than urban sites, rural sites can be 

affected by transport of O3 or O3 precursors from upwind urban areas and by local anthropogenic 

sources such as motor vehicles, power generation, biomass combustion, or oil and gas 

operations” (U.S. EPA, 2013, section 3.6.2.2). In addition, O3 tends to persist longer in rural 

than in urban areas due to lower rates of chemical scavenging in non-urban environments. At 

higher elevations, increased O3 concentrations can also result from stratospheric intrusions (U.S. 

EPA, 2013, sections 3.4, 3.6.2.2). As a result, O3 concentrations measured in some rural sites 

can be higher than those measured in nearby urban areas (U.S. EPA, 2013, section 3.6.2.2). 

These known differences between urban and rural sites suggest that there is the potential for 8-

hour daily peak O3 concentrations and cumulative, seasonal exposures to not correlate well in 

those areas. However, while these metrics may not be directly correlated, reductions in NOx 

emissions that occur in urban areas to attain primary standards would also have the effect of 

reducing downwind, rural concentrations over the season. 

In addition, as was done in both the 1997 and 2008 reviews, staff has analyzed 

relationships between O3 levels in terms of the current averaging time and form and a W126 

cumulative form, based on recent air quality data. One analysis describes the W126 index values 

and current standard design values at each monitor for two periods: 2001-2003 and 2009-2011 

(e.g., Appendix 2B, Figures 2B-2 and 2B-3). This shows that between the two periods, during 

which broad scale O3 precursor emission reductions occurred, O3 concentrations in terms of both 

metrics were reduced. There is a fairly strong, positive degree of correlation between the two 
6-39 

metrics (Appendix 2B).15 Focusing only on the latter dataset (2009-2011), it can be seen that at 

monitors just meeting the current standard (3-year average fourth-highest daily maximum 8-hour 

average concentration equal to 0.075 ppm), W126 index values (in this case 3-year averages) 

varied from less than 3 ppm-hrs to approximately 20 ppm-hrs (Appendix 2B, Figure 2B-3b). At 

sites with a 3-year average fourth-highest daily maximum 8-hour average concentration at or 

below a potential alternative primary standard level of 70 ppb, 3-year W126 index values were 

above 17 ppm-hrs at no monitors, above 15 ppm-hrs at one monitor, and above 13 ppm-hrs at 8 

monitors. At sites with a 3-year average fourth-highest daily maximum 8-hour average 

concentration at or below a potential alternative primary standard level of 65 ppb, 3-year W126 

index values were above 13 ppm-hrs at no monitors, above 11 ppm-hrs at three monitors, and 

above 7 ppm-hrs at 9 monitors. The majority of these monitoring sites are located in the West 

and Southwest and include the states of Arizona, California, Colorado, Nevada, New Mexico, 

and Utah. At sites with a 3-year average fourth-highest daily maximum 8-hour average 

concentration at or below a potential alternative primary standard level of 60 ppb, 3-year W126 

index values were at or below 7 ppm-hrs at all monitors. 

An additional analysis presents the data for sets of recent 3-year periods back to 2006 – 

2008 and indicates that among the counties with O3 concentrations that met the current standard, 

the number of counties with 3-year W126 index values above 15 ppm-hrs ranges from fewer 

than 10 to 24 (Appendix 2B, Figure 2B-9). In general during this longer period, W126 index 

values above 15 ppm-hrs and meeting the current standard were pre-dominantly in Southwest 

region. As the first analysis in Appendix 2B (for the 2001-2003 and 2009-2011 periods) 

indicates, monitors in the West and Southwest tend to have higher W126 index values relative to 

their design values than do monitors in other regions. This pattern is noteworthy because the 

Southwest region has a less dense monitoring network than regions in the Eastern U.S. (see 

Figure 2-1), so that the extent to which this pattern occurs throughout these regions is uncertain. 

Although single-year W126 index values were not separately analyzed in this analysis of the 

monitor data, it indicates appreciable variation in cumulative, seasonal O3 concentrations among 

monitor locations meeting different levels of a standard of the current form. 

Analyses of the WREA air quality scenarios indicate the potential for O3 precursor 

emission reductions achieving O3 concentrations that just meet different 8-hour standards to 

produce a significant reduction in 3-year W126 index values. For example, for the current 

standard scenario, nearly all adjusted monitors are at or below an estimated 3-year average W126 

index value of 15 ppm-hrs (as summarized in section 5.2.2 and described in U.S. EPA, 2014b, 

 

15 Appendix 2B additionally observes that the program implemented for reducing precursor emissions, 

especially NOx, appears to have been an effective strategy for lowering both design values and W126 index values. 
6-40 

Table 4-1). Those monitors above 15 ppm-hrs would be limited to large urban areas in the 

southwestern U.S. (i.e., Phoenix, Los Angeles and Denver). When meeting a 4th highest 8-hour 

average scenario of 70 ppb averaged across 3 years, nearly all monitors in the U.S. would meet a 

3-year W126 index value of 11 ppm-hrs, though some monitors in the southwest would remain 

between 11 and 15 ppm-hrs. At 65 ppb, all locations are at or below 11 ppm-hrs. Thus, similar 

to the monitoring analysis, the modeling analysis generally indicates reductions in W126 levels 

with reduced O3 concentrations in terms of the current standard averaging time and form. This 

suggests that depending on the level for a standard of the current averaging time and form, a 

degree of welfare protection may be afforded. The extent to which such protection provides 

adequate public welfare protection additionally depends on the level of protection identified by 

the Administrator as requisite to protect the public welfare from any known or anticipated 

adverse effects. In so noting, however, we recognize the importance of also considering 

uncertainties in both the model-based adjustment analyses and those based on monitoring data. 

These uncertainties, including those related to monitor coverage, the extent to which recent data 

can be expected to describe future relationships, and modeling approaches16, among others, 

should be kept in mind when assessing the strength of this apparent relationship. 

6.5 CASAC ADVICE 

In our consideration of potential alternative standards, in addition to the evidence-based, 

risk/exposure-based, and air quality information discussed above, we also consider the advice 

and recommendations of CASAC in EPA’s proposed 2010 reconsideration of the 2008 decision, 

as well as comments received in the current review, in the context of its review of the ISA, and 

the WREA and PA. Some of this advice on specific aspects of the evidence and exposure/risk 

information has already been discussed in the relevant sections above. This section specifically 

considers CASAC’s scientific advice on the appropriate form, averaging times and level(s) 

associated with a secondary standard and other related science and policy advice. We have 

additionally considered public comments received to date, some of which have suggested a lack 

of new information to support a distinct secondary standard and others that urge the 

consideration of a secondary standard with a cumulative seasonal form using the W126 metric 

and a level within the range of 7 to 15 ppm-hrs.17 

 

16 One uncertainty associated with the modeling approach, as noted in Chapter 5, relates to the lowering of 

the highest monitored values as a result of the application of the interpolation method used to estimate W126 index 

values at the centroid of every 12 X 12 km2

 grid resolution, rather than only at the exact location of a monitor. 17 Public comment received thus far in this review are in the docket EPA–HQ–OAR-2008-0699, accessible 

at www.regulations.gov. 
6-41 

In response to the EPA’s solicitation of CASAC’s advice on the Agency’s proposed 

rulemaking as part of the reconsideration,18 CASAC conveyed their support for a secondary 

standard distinct from the primary standard, stating that it “also supports EPA’s secondary ozone 

standard as proposed: a new cumulative, seasonal standard expressed as an annual index of the 

sum of weighted hourly concentrations (i.e., the W126 form), cumulated over 12 hours per day 

(8am to 8pm) during the consecutive 3-month period within the ozone season with the maximum 

index value, set as a level within the range of 7 to [1]5 ppm-hours. This W126 metric can be 

supported as an appropriate option for relating ozone exposure to vegetation responses, such as 

visible foliar injury and reductions in plant growth. We found the Agency’s reasoning ... to be 

supported by the extensive scientific evidence considered in the last review cycle. In choosing 

the W126 form for the secondary standard, the Agency acknowledges the distinction between the 

effects of acute exposures to ozone on human health and the effects of chronic ozone exposures 

on welfare, namely that vegetation effects are more dependent on the cumulative exposure to, 

and uptake of, ozone over the course of the entire growing season (defined to be a minimum of at 

least three months). In this proposal, the Agency is responding to the clear need for a secondary 

standard that is different from the primary standard in averaging time, level and form” (Samet, 

2010, p. i-ii). 

In advice offered in the current review, which considers an updated scientific and 

technical record since the 2008 rulemaking, the CASAC reiterated its earlier conclusions 

regarding the appropriate form and averaging times for a secondary O3 NAAQS at several points 

in its letter to the Administrator. In stating the basis for its conclusion, CASAC notes that “[i]n 

reaching its scientific judgment regarding the indicator, form, summation time, and range of 

levels for a revised secondary standard, the CASAC has focused on the scientific evidence for 

the identification of the kind and extent of adverse effects on public welfare” (Frey, 2014a, p. 

iii), and further that “[t]hese recommendations are based on scientific evidence of adverse effect 

associated with the presence of ozone in ambient air” (Frey, 2014a, p. 15). On this basis, 

CASAC reached its conclusions on the appropriate form for the secondary standard stating “[t]he 

CASAC supports the scientific conclusion in the Second Draft PA that the current secondary 

standard is not adequate to protect against current and anticipated welfare effects of ozone on 

vegetation. We recommend retaining the current indicator (ozone) but establishing a revised 

form of the secondary standard to be the biologically relevant W126 index accumulated over a 

 

18 The reconsideration proposal included a proposed new cumulative, seasonal secondary standard, 

expressed as an index of the annual sum of weighted hourly concentrations (the W126 index), cumulated over 12 

hours per day during the consecutive 3-month period within the O3 season with the maximum index value, averaged 

over three years, set within a range of 7 to 15 ppm-hrs (75 FR 3027). 
6-42 

12-hour period (8 a.m. – 8 p.m.) over the 3-month summation period of a single year resulting in 

the maximum value of W126” (Frey, 2014a, p. iii).

In addition, we take note of the scientific advice provided by CASAC regarding its 

scientific judgments regarding appropriate target benchmarks of protection and the range of 

W126 index values that in its scientific judgment provides appropriate protection for these 

benchmarks. CASAC states that “[a] 2% biomass loss is an appropriate scientifically based 

value to consider as a benchmark of adverse impact for long-lived perennial species such as 

trees, because effects are cumulative over multiple years” and “[c]rop loss appears to be less 

sensitive than these other indicators, largely because of the CASAC judgment that a 5% yield 

loss represents an adverse impact, and in part due to more opportunities to alter management of 

annual crops” (Frey, 2014a, p. 14). 

Given these benchmarks, CASAC provided further advice regarding an appropriate range 

of W126 levels that it considered appropriately protective. Specifically, “[t]he CASAC 

recommends that the level associated with this form be within the range of 7 ppm-hrs to 15 ppmhrs to protect against current and anticipated welfare effects of ozone. The CASAC does not 

support a level higher than 15 ppm-hrs. For example, at 17 ppm-hrs, the median tree species has 

6% relative biomass loss, and the median crop species has over 5% yield loss. These levels are 

unacceptably high” (Frey, 2014a, p. iii)19. CASAC further noted that “[w]ith compounding over 

the harvest cycle or life span of these species, this will result in considerably greater cumulative 

RBL as discussed above. For the more sensitive tree seedlings, a value closer to the lower end of 

the range (7 ppm-hrs) would be more appropriate. The level of 7 ppm-hrs is the only level 

analyzed for which the relative biomass loss for the median tree species is less than or equal to 2 

percent. At 7 ppm-hrs, 7 of the 12 analyzed species have relative biomass loss of less than 2%” 

(Frey, 2014a, p. 14). 

CASAC further noted that “the correlative similarity between the current standard and a 

level of the W126 index of 15 ppm-hrs must not be interpreted to mean that just meeting the 

current standard is equivalent to just meeting a W126 level of 15 ppm-hrs. Most of the analyses 

found effects below 15 ppm-hrs (many at 10 or even 7 ppm-hrs)” (Frey, 2014a, p. 12). 

CASAC also recognized that there were policy choices left to the Administrator with 

respect to determining an appropriate level of protection. In so doing “[t]he CASAC 

acknowledges that the choice of a level within the range recommended based on scientific 

evidence is a policy judgment under the statutory mandate of the Clean Air Act. Specifically, the 

 

19 As noted in Section 6.3, the numbers for RBL for the median tree species have been updated between the 

second and final PA to deemphasize cottonwood, based on staff’s understanding of CASAC advice in that regard. 

We note that CASAC advice based on what is shown in Table 6-1 is no longer consistent in some cases with the 

revised table, and in particular with regard to median tree seedling RBL values. 
6-43 

Clean Air Act grants discretion to the Administrator to specify a standard that is ‘requisite to 

protect the public welfare from any known or anticipated adverse effects associated with the 

presence of [the] pollutant in the ambient air’... (Frey, 2014a, p. iii). In addition, CASAC also 

offered its policy advice regarding selection of an appropriate level within its scientifically 

recommended range, stating that “[a]s a policy recommendation, separate from its advice above 

regarding scientific findings, the CASAC advises that a level of 15 ppm-hrs for the highest 3-

month sum in a single year is requisite to protect crop yield loss, but that lower levels provide 

additional protection against crop yield loss. Furthermore, there are specific economically 

significant crops, such as soybeans, that may not be protected at 15 ppm-hrs but would be 

protected at lower levels. A level below 10 ppm-hrs is required to reduce foliar injury. A level of 

7 ppm-hrs is protective of relative biomass loss for trees and offers additional protection against 

crop yield loss and foliar injury. Therefore, 7 ppm-hrs is protective of ecosystem services. Thus, 

lower levels within the recommended range offer a greater degree of protection of more 

endpoints than do higher levels within the range” (Frey, 2014a, p. iii). 

Additionally, in regard to the 3-year average option discussed in the second draft PA, 

CASAC thus notes that “[i]f, as a policy matter, the Administrator prefers to base the secondary 

standard on a three-year averaging period for the purpose of program stability, then the level of 

the standard should be revised downward such that the level for the highest three-month 

summation in any given year of the three-year period would not exceed the scientifically 

recommended range of 7 ppm-hrs to 15 ppm-hrs. ...The final Policy Assessment should quantify 

the ratio of the three-year average of the highest three-month summations in each year to the 

highest three-month summation in the highest year. This ratio should be used to determine what 

downward adjustment from the three-month summation in one year recommended here is needed 

if a three-year form is selected” (Frey, 2014a, pp. iii and iv).20

Finally, we note that in commenting on the significance of the uncertainties associated 

with the evidence and exposure and risk analyses that remain, CASAC concludes that “[w]hile 

these scientific research priorities will enhance future scientific reviews of the ozone primary and 

secondary standards, we also make clear that there is sufficient scientific evidence, and sufficient 

confidence in the available research results, to support the advice we have given above for this 

review cycle of the primary and secondary standards” (Frey, 2014a, p. iv). 

 

20 See Section 6.4 and Chapter 2 for more discussion on the relationship between one-year and three-year 

average W126 index values. 
6-44 

6.6 STAFF CONCLUSIONS ON ALTERNATIVE STANDARDS 

Staff’s consideration of alternative secondary O3 standards builds on our conclusion from 

section 5.7 above that the body of evidence, in combination with the results of the WREA 

analyses, calls into question the adequacy of the current secondary standard and provides support 

for consideration of alternative standards. In sections 6.1 to 6.3 above, we consider how the 

currently available scientific evidence and exposure/risk information informs staff conclusions 

regarding the basic elements of the NAAQS: indicator (6.1), form and averaging time (6.2), and 

level (6.3). In so doing, we consider both the information available at the time of the last review 

and information newly available since the last review that has been critically analyzed and 

characterized in the 2013 ISA. As an initial matter, with regard to the indicator, we conclude 

that based on the available science it is still appropriate to continue to use measurements of O3 in 

accordance with federal reference methods as the indicator to address effects associated with 

exposure to ambient O3 alone or in combination with related photochemical oxidants. 

 In considering alternative standards, staff has considered the available body of evidence 

as comprehensively assessed in the ISA, the risk and exposure information presented in the 

WREA, and CASAC advice and public comment in this review with regard to support for 

consideration of options that are different from the current standard, as articulated by the 

following overarching question: 

 To what extent does the currently available scientific evidence- and exposure/riskbased information, as reflected in the ISA and WREA, support consideration of 

alternatives to the current O3 standard to provide increased protection from 

ambient O3 exposures? 

In considering potential forms alternative to that of the current standard, we note that the 

form for the current secondary standard is the 4th highest daily maximum 8-hour average, 

averaged over three years. As discussed in chapter 5 and section 6.2 above, the longstanding 

evidence regarding the fundamental aspects of O3 exposure that are directly responsible for 

inducing vegetation response indicates that plant response to O3 is driven by the cumulative 

exposure to O3 during the growing season (U.S. EPA, 2013, section 2.6.6.1). This cumulative 

exposure depends on both the total duration of the exposure (from repeated O3 episodes) and the 

concentrations of those exposures (higher concentrations having a disproportionate impact as 

compared to lower concentrations). On the basis of this longstanding and extensive evidence, 

the ISA concludes that exposure indices that cumulate and differentially weight the higher hourly 

average concentrations over a season and also include the mid-level values offer the most 

scientifically defensible approach for use in developing response functions and in defining 

indices for vegetation protection (U.S. EPA, 2013, section 2.6.6.1). 
6-45 

CASAC advice in the 2008 review and on the 2010 proposed reconsideration also 

recognized that the nature of the exposures relevant to vegetation response is well described by a 

cumulative seasonal form and has supported the use of such a form for a secondary O3 standard 

(Henderson, 2006; Samet, 2010). The current CASAC O3 Panel has expressed similar views. 

We also note that on the basis of the evidence and exposure/risk information available in the two 

previous reviews, and in consideration of CASAC advice, the Administrator has recognized the 

importance of protecting vegetation from cumulative, seasonal exposures and proposed such a 

form as an appropriate, reasonable policy option (61 FR 65741-44; 72 FR 37899-905; 75 FR 

3012-3027). 

Thus, in considering alternative forms of the standard we conclude that it is reasonable 

and appropriate to consider a cumulative, concentration-weighted form to provide protection 

against cumulative, seasonal exposures to O3 that are known or anticipated to harm sensitive 

vegetation or ecosystems. Such a form is specifically designed to focus on the kind of O3

exposures that have been shown to cause harm to vegetation and would have a distinct advantage 

over the form of the current standard in characterizing air quality conditions potentially of 

concern for vegetation and in more directly demonstrating that the desired degree of protection 

against those conditions was being achieved. 

In considering the appropriate index for a cumulative seasonal form, we recognize that a 

number of different cumulative concentration weighted indices have been developed and have 

been evaluated in the scientific literature and in past NAAQS reviews in terms of their ability to 

predict vegetation response and their usefulness in the NAAQS context (U.S. EPA, 2006, pp. 9-

11 to 9-15 and pp. AX9-159 to AX9-187; U.S. EPA, 2007, pp. 7-15/16). While these various 

forms have different strengths and limitations, as noted in the ISA (U.S. EPA, 2013, section 9.5), 

the W126 index21 has some important advantages over other non-sigmoidally weighted 

cumulative indices. For example, given the lack of a discernible threshold for vegetation effects 

in general, we recognize the fact that the W126 metric does not have a cut-off in its weighting 

scheme (down to about 30 ppb below which the weighting factor is effectively zero), such that it 

includes consideration of potentially damaging lower O3 concentrations. Additionally, the W126 

metric adds increasing weight to hourly concentrations from about 40 ppb to about 100 ppb, an 

important feature because “as hourly concentrations become higher, they become increasingly 

likely to overwhelm plant defenses and are known to be more detrimental to vegetation” (U.S. 

 

21 The W126 is a non-threshold approach described as the sigmoidally weighted sum of all hourly O3

concentrations observed during a specified diurnal and seasonal exposure period, where each hourly O3

concentration is given a weight that increases from 0 to 1 with increasing concentration (Lefohn et al, 1988; Lefohn 

and Runeckles, 1987; U.S. EPA, 2013, section 9.5.2). 
6-46 

EPA, 2013, p. 9-104). We additionally take note of CASAC advice in the 2008 review and on 

the 2010 proposed reconsideration recommending the use of the W126 index for a cumulative 

seasonal form for a secondary O3 standard (Henderson, 2006; Samet, 2010). Similarly, the 

current CASAC O3 Panel has indicated that a focus on a W126 form is appropriate (Frey, 2014a, 

p. iii). Therefore, on the basis of the strength of the evidence and advice from CASAC, we 

conclude that the W126 index is the most appropriate cumulative seasonal form to consider in 

the context of the secondary O3 NAAQS review. 

We next turn to the exposure periods – diurnal and seasonal – over which the W126 

index would be summed in any given year. As discussed in section 6.2 above, the currently 

available information continues to provide support for a definition of the diurnal period of 

interest as the 12-hour period from 8:00 am to 8:00 pm (U.S. EPA, 2013, section 9.5.3). In prior 

reviews, the EPA has identified the 12-hour period from 8:00 am to 8:00 pm as appropriately 

capturing the diurnal window with most relevance to the photosynthetic process (72 FR 37900; 

75 FR 3013), and CASAC has generally supported the 12-hour daylight period (Henderson, 

2006, 2007). In light of the continued support in the evidence base and no evidence on this issue 

differing from that in previous reviews, we again conclude that it is appropriate to use the 12-

hour period from 8:00 am to 8:00 pm to cumulate daily O3 exposures. On this basis, we 

conclude that the 12-hour daylight window (8:00 am to 8:00 pm) represents the portion of the 

diurnal exposure period that is most relevant to predicting or inducing plant effects related to 

photosynthesis and growth and thus is an appropriate diurnal period to use in conjunction with a 

W126 cumulative metric. 

With regard to a seasonal period of interest, the current evidence base continues to 

provide support for a seasonal period with a minimum duration of three months (U.S. EPA, 

2013, section 9.5.3). We note that a plant is vulnerable to O3 pollution as long as it has foliage 

and is physiologically active (U.S. EPA, 2013, section 9.5.3, p. 9-112), i.e., during its growing 

season. The exposure periods used in studies of O3 effects on vegetation reflect this 

understanding and typically focus on study periods of 3-6 months. Included in the currently 

available evidence is a new analysis that compared 3- and 6-month maximum W126 index 

values for over 1,200 AQS and CASTNET EPA monitoring sites for the years 2008-2009 that 

found that the two accumulation periods were highly correlated (U.S. EPA, 2013, section 9.5.3, 

Figure 9-13). Thus, although we recognize that the selection of a single seasonal time period 

over which to cumulate O3 exposures for a national standard necessarily represents a balance of 

factors, given the significant variability in growth patterns and lengths of growing season among 

vegetative species growing within the U.S., we conclude it is appropriate to identify the seasonal 

W126 index value as that derived from the consecutive 3-month period within the O3 season 

with the highest W126 index value. We note that such a 3-month exposure period was also 
6-47 

supported by CASAC in advice provided during the last review and the 2010 proposed 

reconsideration (Henderson, 2006; Samet, 2010). 

With regard to form, we additionally consider the period of time over which a cumulative 

seasonal W126-based standard should be evaluated. In so doing, we have considered the support 

for both a single year form and a form averaged over three years (section 6.2). We note 

comments from CASAC on this matter, in particular their comment in the current review that 

“[t]he CASAC does not recommend the use of a three-year averaging period for the secondary 

standard. We favor a single-year period for determining the highest three-month summation 

which will provide more protection for annual crops and for the anticipated cumulative effects on 

perennial species. The scientific analyses considered in this review, and the evidence upon which 

they are based, are from single-year results” (Frey, 2014a, p. iii). 

We recognize that there are a number of O3-induced effects that have the potential for 

public welfare significance within the annual timeframe. These effects mainly include reduced 

crop yields and visible foliar injury, as noted in section 6.2 above. There are uncertainties 

associated with these effects that make it difficult to determine the degree of annual protection 

needed to protect the public welfare from any known or anticipated adverse effects. There are 

also annual effects in perennial species that may result from a single year exposure and can be 

“carried over” into the subsequent year where they affect growth and reproduction (U.S. EPA, 

2013, pp. 9-43 to 9-44 and p. 9-86). When such annual effects due to elevated O3 exposures 

occur over multiple years, they have the further potential to be compounded, increasing the 

potential for effects at larger scales (e.g., population, ecosystem), including effects on associated 

services that may be of significance to the public welfare. These ecosystem services effects can 

include alteration of below-ground biogeochemical cycles, and alteration of both above- and 

below-ground terrestrial community composition and terrestrial ecosystem water cycling (U.S. 

EPA, 2013, Table 9-19) and reductions in productivity and carbon sequestration in terrestrial 

ecosystems. We additionally note that multiple consecutive years of critical O3 exposures might 

be expected to result in larger impacts on forested areas (e.g., increased susceptibility to other 

stressors such as insect pests, disease, co-occurring pollutants and harsh weather) than 

intermittent occurrences of such exposures due to the potential for compounding or carry-over 

effects on tree growth. 

Given the above, we conclude that the public welfare significance of the effects that can 

occur as a result of three-year O3 exposures are potentially greater than those associated with a 

single year of such exposure. Thus, to the extent that the focus for public welfare protection to 

be afforded by the secondary O3 standard is on long-term effects that occur in sensitive tree 

species in natural forested ecosystems, including federally protected areas such as Class I areas 

or on lands set aside by States, Tribes and public interest groups to provide similar benefits to the 
6-48 

public welfare, a standard with a form that evaluates the cumulative seasonal index across 

multiple years might be considered to provide a more appropriate match to the nature of O3-

related effects on vegetation upon which the secondary O3 standard is focused. In considering 

such forms, we focus on one that averages the W126 index values across three years, as 

discussed in section 6.2 above. 

In addition to the vegetation effects considerations described above, there are other 

policy-relevant factors that can be useful to consider. For example, under a standard with a 

single year form, a monitor may be judged to meet the standard based on a single year of data, 

while under a standard with a form requiring evaluation over a multi-year period, a monitor is 

not judged to have met the standard until a complete multi-year record is available. For a W126-

based potential standard, the multi-year form identified for consideration in the last review was 

the average cumulative seasonal metric over three consecutive years (75 FR 3027). Such a 

multi-year form remains appropriate to consider to provide stability to an alternative secondary 

standard, just as the multi-year form provides for the current standard (average over three years 

of annual fourth-highest daily maximum 8-hour average O3 concentrations).22 In considering the 

issue of stability in the context of such a form, we first note the inter-annual variability of 

seasonal W126 index, which is not unexpected given the logistic weighting function and also 

inter-annual variability in meteorological conditions that contribute to O3 formation (see 

Appendix 2C). The staff analysis in Appendix 2C describes the variability in annual W126 

index values in relation to variability in the 3-year average, which indicates that a standard based 

on an annual W126 index would be expected to have a lower degree of year-to-year stability 

relative to a standard based on a form that averages seasonal indices across three consecutive 

years. A more stable standard can be expected to contribute to greater public welfare protection 

by limiting year-to-year disruptions in ongoing control programs that would occur if an area was 

frequently shifting in and out of attainment due to extreme year-to-year variations in 

meteorological conditions. This greater stability in air quality management programs thus 

facilitates achievement of the protection intended by a standard. In light of this relationship, we 

conclude that a 3-year average form has the desirable feature of providing greater stability in air 

quality management programs and thus facilitating the achievement of the protection intended by 

a standard. Thus, we recognize the public welfare benefits of having a standard of a 3-year 

average form. 

Thus, to the extent that the greater emphasis is placed on protecting against effects 

associated with multi-year exposures and maintaining more year-to-year stability of public 

 

22 See ATA III, 283 F. 3d at 374-75 (recognizing programmatic stability as a legitimate consideration in the 

NAAQS standard-setting process). 
6-49 

welfare protection, we conclude that it is appropriate to consider a secondary standard form that 

averages the seasonal W126 index values across three consecutive years. We conclude that such 

a form might be appropriate for a standard intended to achieve the desired level of protection 

from longer-term effects, including those associated with potential compounding. Further, such 

a form might be concluded to contribute to greater stability in air quality management programs, 

and thus, greater effectiveness in achieving the desired level of public welfare protection, than 

that that might result from a single year form. 

Turning to consideration of an appropriate range of levels for a W126-based standard, we 

first note that our general approach to informing these judgments recognizes that the available 

evidence demonstrates a range of O3 sensitivity across studied plant species and documents an 

array of O3-induced effects that extend from lower to higher levels of biological organization. 

These effects range from those affecting cell processes and individual plant leaves to effects on 

the physiology of whole plants, species effects and effects on plant communities to effects on 

related ecosystem processes and services. Given this evidence, it is not possible to generalize 

across all studied species regarding which cumulative exposures are of greatest concern, as this 

can vary by situation due to differences in exposed species sensitivity, the importance of the 

observed or predicted O3-induced effect, the role that the species plays in the ecosystem, the 

intended use of the affected species and its associated ecosystem and services, the presence of 

other co-occurring predisposing or mitigating factors, and associated uncertainties and 

limitations. At the same time, the evidence also demonstrates that though effects of concern can 

occur at very low exposures in sensitive species, at higher cumulative exposures those effects 

would likely occur at a greater magnitude and/or higher levels of biological organization and 

additional species would likely be impacted. It is important to note, however, that due to the 

variability in the importance of the associated ecosystem services provided by different species 

at different exposures and in different locations, as well as differences in associated uncertainties 

and limitations, that, in addition to the magnitude of the ambient concentrations, both the species 

present and their public welfare significance are essential considerations in drawing conclusions 

regarding the significance or magnitude of public welfare impact. 

Therefore, in developing conclusions in this PA, we take note of the complexity of 

judgments to be made by the Administrator regarding the adversity of known and anticipated 

effects to the public welfare and are mindful that the Administrator’s ultimate judgments on the 

secondary standard will, as appropriate, reflect an interpretation of the available scientific 

evidence and exposure/risk information that neither overstates nor understates the strengths and 

limitations of that evidence and information. 

As described above in section 5.1, we employ a paradigm to assist in putting the available 

science and exposure/risk information into the public welfare context. This paradigm has 
6-50 

evolved over the course of the O3 NAAQS reviews and has also been informed by similar 

constructs developed for other secondary NAAQS reviews. As discussed in Section 5.1, this 

paradigm recognizes that the significance to the public welfare of O3-induced effects on sensitive 

vegetation growing within the U.S. can vary depending on the nature of the effect, the intended 

use of the sensitive plants or ecosystems, and the types of environments in which the sensitive 

vegetation and ecosystems are located. Accordingly, any given O3-related effect on vegetation 

and ecosystems (e.g., biomass loss, crop yield loss, visible foliar injury) may be judged to have a 

different degree of impact on or significance to the public welfare depending, for example, on 

whether that effect occurs in a Class I area, a city park, or commercial cropland. In the last 

review, the Administrator placed the highest priority and significance on vegetation and 

ecosystem effects to sensitive species that are known to or are likely to occur in federally 

protected areas such as national parks and other Class I areas, or on lands set aside by states, 

tribes and public interest groups to provide similar benefits to the public welfare (75 FR 3023-

24; 73 FR 16496), recognizing that effects occurring in such areas would likely have the highest 

potential for being classified as adverse to the public welfare, due to the expectation that these 

areas need to be maintained in pristine or near pristine conditions to ensure their intended use is 

met. This approach also includes consideration of impacts to ecosystem goods and services. 

Although ecosystem services were not explicitly considered in the Administrator’s decision in 

the last review, they were explicitly recognized as an important category of public welfare 

effects and they have an obvious relationship to consideration of intended use (73 FR 16492). In 

employing this approach, we note the support for it provided by CASAC advice in this review 

(Frey, 2014a). 

In considering potential levels for an alternative standard based on the W126 metric, we 

focus the discussion primarily on: 1) impacts on tree growth, productivity and carbon storage; 2) 

crop yield loss; and 3) visible foliar injury. With respect to tree growth, we find it useful to 

consider the summary of relative biomass loss estimates in Table 6-1 above and the WREA 

risk/exposure estimates discussed in Section 6.3 and Appendix 6F. In Table 6-1, we take note of 

the different index value estimates with regard to the number of studied species below different 

response benchmarks, as well as with regard to the median response. We additionally consider 

the WREA estimates regarding: (1) percent of assessed geographic area exceeding 2% weighted 

relative biomass (Table 6-2); (2) number of assessed Class I areas with tree seedling weighted 

relative biomass loss estimates above 2% (Table 6-3); and (3) the percent median biomass loss 

across counties for different air quality scenarios (Table 5-5). Further, we note other WREA 

estimates for effects on ecosystem services related to public welfare, such as carbon 

sequestration and air pollutant removal. With respect to crop yield loss, we note the summary of 

crop yield loss estimates in Table 6-1 and the WREA risk/exposure estimates discussed in 
6-51 

Section 6.3 and Appendix 6F, which include individual species and median response. We also 

note information available on visible foliar damage to species occurring in natural settings, such 

as federal Class I areas, and the analyses in the WREA evaluating biosite data and several 

benchmarks of injury as summarized in section 5.4.2. 

In focusing on trees and their associated ecosystem services, we first note that the studied 

tree species vary widely in their sensitivity to O3-induced relative biomass loss. We thus find it 

informative to consider both median species values and individual species responses and RBL 

over the same W126 range. We note CASAC’s advice regarding RBL levels, specifically their 

emphasis on a benchmark of median relative tree biomass loss at or below 2% and their view 

that a 6% median relative biomass loss is “unacceptably high”. From Table 6-1 we see that 

median tree species biomass loss is at or below 2% only at the lowest W126 level assessed, 7 

ppm-hrs. As the W126 level is incrementally increased, median RBL also increases 

incrementally, so that at W126 index values of 9, 11, 13, 15, 17 and 19, the median RBL 

increases to 2.4%, 3.1%, 3.8%, 4.5%, 5.3% and 6.0%, respectively. Thus over the W126 range 

of 7 – 17 ppm-hrs, median species biomass loss ranges from approximately 2% to approximately 

5%. 

 We next take note of the number of individual species’ RBLs that fall below those same 

benchmarks assessed for median species values. We also note the value of additionally 

characterizing the RBL estimates in comparison to higher loss levels such as 10% or 15%, 

especially for individual tree species. Based on Figure 5-1 (B) in Chapter 5, and as shown in 

Table 6-1, for W126 values at or below 17 ppm-hrs, the RBLs for each of 5 species is less than 

2%. Thus, over the full range of alternative levels considered, the same level of protection 

relative to the 2% benchmark is achieved for these species. We therefore turn our attention to 

the remaining 6 studied species to see if additional information might be available to help inform 

consideration of an appropriate degree of protection. Specifically, we consider the RBL 

information available for the other species (i.e., eastern white pine, aspen, tulip poplar, 

ponderosa pine, red alder, and black cherry) to further inform our evaluation of the additional 

protection that potentially could be achieved at different W126 levels within the range identified. 

We note that, for W126 levels of 17 to 7 ppm-hrs, biomass loss decreases for these individual 

species with decreasing W126 levels such that at the W126 level of 17 ppm-hrs, five species 

have RBL above 6% while at the W126 level of 7 ppm-hrs, one species (black cherry) has an 

RBL above 6%. Taken together with the more tolerant species, the proportion of the studied tree 

species with RBLs below 6% are 6/11, 7/11, 8/11, and 10/11 at W126 index values of 17, 15, 

13, and 11 ppm-hrs, respectively. 

 In consideration of other benchmark levels, 9/11 studied tree species have a predicted 

RBL below 10% at the W126 level of 17 ppm-hrs, while 10/11 species have a predicted RBL 
6-52 

below 10% for W126 levels of 15 to 7 ppm-hrs. In addition, 10/11 studied tree species have a 

predicted RBL below 15% for W126 levels of 17 to 7 ppm-hrs. We note that black cherry, the 

most sensitive of the 11 species, has RBLs ranging from approximately 36% at W126 index 

value of 17 down to approximately 17% at the W126 index value of 7 ppm-hrs. Thus, the 

predicted RBL for black cherry remains above 15% for W126 levels of 17 to 7 ppm-hrs, and it is 

not clear to what extent those predicted RBL values might inform consideration of the level of 

protection achieved for different W126 exposures within this range (Table 6-1; U.S. EPA, 

2014b, section 6.2, Appendix 6A). 

To further inform this issue, the WREA also characterizes the number of counties where 

the median RBLs were greater than 2% (U.S. EPA, 2014b, Table 6-7), as shown in Table 5-5. 

When air quality is adjusted to the current standard, 8% of the counties have median RBLs 

greater than 2%. That proportion drops to 7% for air quality adjusted to just meet a 3-year 

average W126 level of 15 ppm-hrs and to 6% for air quality adjusted to just meet a 3-year 

average W126 level of 7 ppm-hrs. Of the 239 counties (8% of counties) estimated to have a 

median RBL above 2% when meeting the current standard, 203 of those counties have a RBL 

greater than 2% because of the presence of black cherry. Thus, as also discussed above in 

Section 6.2, given the large magnitude of estimated RBL for black cherry over the entire range 

assessed, it is not clear to what extent the information for black cherry informs consideration of 

the overall level of protection achieved across the identified range. 

In considering the potential magnitude of the ecosystem impact of reduced biomass in 

trees, we focus on the WREA estimates of weighted RBL for the W126 air quality scenarios 

(U.S. EPA, 2014b, section 6.8), focusing particularly on impacts in Class I areas. For the current 

standard and the three W126 scenarios (15 ppm-hr, 11 ppm-hr, and 7 ppm-hr), the percent of 

total national land-area having weighted RBL greater than 2% was 0.2%, 0.2%, 0.1% and 

<0.1%, respectively (Table 6-2; U.S. EPA 2014, Table 6-25). In addition, the WREA estimates 

indicate weighted RBL greater than 2% in 1-2 of 145 assessed nationally protected Class I areas 

for the current standard and all three W126 scenarios. To the extent that emphasis is given to 

such estimates for nationally protected Class I areas and for appreciable percentages of forested 

areas nationwide, a W126 index value extending up to 17 ppm23 may be appropriate to consider. 

The WREA provides qualitative and semi-quantitative information regarding the types 

and potential magnitude of O3 impacts on ecosystem services. In noting the potential ecosystem 

 

23 While the WREA analyses did not include an air quality scenario for 17 ppm-hrs, the data suggests that, 

to adjust air quality to meet a W126 index value of 17 ppm-hrs, additional emissions reductions would have been 

needed relative to just meeting the current standard. Therefore, because the air quality scenarios for meeting the 

current standard and meeting a W126 index value of 15 ppm-hrs both indicate weighted relative biomass loss less 

than or equal to 2% in 143 of 145 assessed nationally protected Class I areas, the same would be true for an air 

quality scenario for just meeting a W126 index value of 17 ppm-hrs. 
6-53 

services benefits related to reductions in tree biomass loss resulting from just meeting potential 

alternative W126-based standards, we recognize, in particular, that impacts on climate regulation 

can reasonably be concluded to be potentially significant from a public welfare perspective and 

carbon sequestration has been identified as a potentially important tool for managing 

anthropogenic impacts on climate. The WREA estimates the potential increase in carbon storage 

that potentially could occur for different air quality scenarios (U.S. EPA 2014, section 6.6.1). 

Comparisons of the W126 scenarios to the current standard scenario with regard to carbon 

sequestration estimates do not indicate an appreciable difference for the W126 scenario of 15 

ppm-hrs beyond that achieved by just meeting the current standard. The majority of the 

enhanced carbon sequestration potential in forests over time is predicted to occur for the 

alternative W126 scenarios of 11 and 7 ppm-hrs. Over 30 years, the current standard scenario 

projection is 89,184 million metric tons of CO2 equivalents (MMtCO2e).24 The WREA estimates 

additional sequestration potential of 13, 593 and 1,600 MMtCO2e for the W126 scenarios of 15, 

11 and 7 ppm-hrs, respectively, as compared to the current standard (U.S. EPA 2014, Table 6-

18). We additionally consider the WREA estimates for five urban areas of how reduced growth 

of O3-sensitive trees in urban forests may affect air pollutant removal (U.S. EPA, 2014b, sections 

6.6.2 and 6.7 and Appendix 6D). Estimates for all five case study areas indicate increased 

pollutant removal from the recent conditions to just meeting the current standard, with much 

smaller differences between the current standard and the three W126 scenarios (Table 6-5). 

However, we additionally take note of significant uncertainties and limitations associated with 

WREA estimates related to carbon sequestration and air pollution removal. Thus, we note that 

an identification of the requisite protection for forest trees and their associated ecosystem 

services would likely involve policy judgments regarding the appropriate weight to place on 

potential impacts to the public welfare with respect to estimated effects on the ecosystem 

services of carbon storage and urban air pollution removal associated with tree growth, as well as 

on the uncertainties associated with this information. 

With respect to crops, we focus on the 10 robust E-R functions (barley, lettuce, field 

corn, grain sorghum, peanut, winter wheat, cotton, soybean, potato and kidney bean) described in 

the ISA and additionally analyzed in the WREA (Figure 5-4). We also note CASAC’s advice 

regarding a recommended target benchmark protection level of 5% for median crop relative yield 

loss (RYL) and that, as shown in Table 6-1, W126 index values ranging from 7 to 17 ppm-hrs 

 

24 1 million metric tons of carbon dioxide equivalents (MMtCO2e) is equivalent to 208,000 passenger 

vehicles or the electricity to run 138,000 homes for 1 year as calculated by the EPA Greenhouse Gas Equivalencies 

Calculator (updated September 2013 and available at http://www.epa.gov/cleanenergy/energyresources/calculator.html). 
6-54 

are estimated to have median crop RYL of less than or equal to approximately 5%. Given this, it 

is not clear to what extent this information informs the selection of an appropriate level. 

When individual species are considered over this same range, the proportion of crops 

protected varies from 5/10, 6/10, 6/10, 9/10, 10/10, and 10/10 at the W126 levels of 17, 15, 13, 

11, 9, and 7 ppm-hrs. To the extent a given species is judged as having particular importance to 

the public welfare, breaking the information down by species can be helpful. For example, less 

than 5% yield loss was estimated for soybeans at the W126 index value of 12 ppm-hrs (U.S. 

EPA 2014, Figure 6-3). Four of the studied crop species (i.e., barley, lettuce, field corn, and 

grain sorghum) are more tolerant, with RYL under 1% over the W126 range from 7 to 17 ppmhrs. Peanut also remained under 4% RYL over the same W126 range. Other species differed 

regarding the W126 level at which RYL reached or fell below 5%. Specifically, for winter 

wheat, cotton, soybean, kidney bean and potato, the relevant W126 index values at which RYLs 

were below 5% are 15, 13, 11, 9 and 7 ppm-hrs. As noted in Chapter 5, and in early discussions 

in this chapter, the significance of these predicted RYLs to the public welfare could be informed 

by the recognition that crops are heavily managed to obtain the desired yield and the potential 

adversity to public welfare from yield reductions in any specific crop in a particular location 

would depend on a number of economic factors, including crop prices, crop substitution, and the 

welfare importance of relative changes in consumer and producer surplus. We also note that 

these crop species would likely receive some protection from a standard set, for example, to 

provide protection against tree biomass loss, such as in areas set aside to be maintained in a more 

pristine condition (75 FR 3024).

Visible foliar injury has been identified by the FLMs as a diagnostic tool for informing 

conclusions regarding potential ozone impacts on potentially sensitive AQRVs (USFS, NPS, 

FWS, 2010), indicating that such O3-induced impacts might be considered to have the potential 

to impact the public welfare in scenic and/or recreational areas during years they occur. We take 

note of the WREA analyses of the nationwide dataset (2006-2010) for USFS/FHM biosites 

described in section 5.4.2 above, including the observation that the proportion of biosites with 

injury varies with soil moisture conditions and O3 W126 index values (U.S. EPA 2014, Chapter 

7, Figure 7-10). These analyses also show that foliar injury incidence increases steeply with 

W126 index values up to approximately 10 ppm-hrs. At W126 index levels greater than that, 

little or no further increase in proportion of sites showing foliar injury occurs. 

With respect to visible foliar injury, we are unaware of any guidance for federal land 

managers regarding at what spatial scale or what degree of severity visible foliar injury is 

sufficient to trigger protective action based on this potential impact on AQRVs. Further, there 

does not appear to be any consensus in the literature regarding severity of foliar injury and risks 

to plant functions or services, and CASAC, while identifying target percent biomass loss and 
6-55 

yield loss benchmarks for tree seedlings and commodity crops, respectively, did not provide a 

similar recommendation for this endpoint. Likewise, as in previous reviews, the ISA notes the 

difficulty in relating visible foliar injury symptoms to other vegetation effects such as individual 

plant growth, stand growth, or ecosystem characteristics (U.S. EPA, 2013, section 9.4.2, p. 9-39) 

and further noted that the full body of evidence indicates that there is wide variability in this 

endpoint, such that although evidence shows visible foliar injury can occur under very low 

cumulative O3 concentrations, “...the degree and extent of visible foliar injury development 

varies from year to year and site to site..., even among co-members of a population exposed to 

similar O3 levels, due to the influence of co-occurring environmental and genetic factors” (U.S. 

EPA 2013, section 9.4.2, p. 9-38). Given this, it is not clear to what extent this information 

informs the selection of an appropriate level. 

On the basis of all the considerations described above, including the evidence and 

exposure/risk analyses, and advice from CASAC, we conclude that a range of W126 index 

values appropriate for the Administrator to consider extends from 7 to 17 ppm-hrs. In so doing, 

however, we note, as recognized above, the role of judgments by the Administrator in such 

decisions. In selecting the range identified here, we primarily consider the evidence- and 

exposure/risk-based information for cumulative seasonal O3 exposures represented by W126 

index values (including those represented by the WREA average W126 scenarios) associated 

with biomass loss in studied tree species, both in and outside areas that have been afforded 

special protections. We note CASAC’s advice that a 6% median RBL is unacceptably high, that 

the 2% median RBL is an important benchmark to consider, that for the lower W126 value of 7 

ppm-hrs that the median tree species biomass loss is at or below 2%, and that for the upper value 

of 17 ppm-hrs the median tree biomass loss is below 6%25. We also note the estimates indicating 

that a W126 level of 17 ppm-hrs reduces the percent of total nationwide land-area having 

weighted RBL greater than 2% to 0.2% (Table 6-2) and the number of Class I areas with 

weighted RBL greater than 2% to 2 of the 145 assessed nationally protected Class I areas. 

We also note that tree biomass loss can be an indicator of more significant ecosystemwide effects which might reasonably be concluded to be significant to public welfare. For 

example, when it occurs over multiple years at a sufficient magnitude, it is linked to an array of 

effects on other ecosystem-level processes, such as nutrient and water cycles, changes in above 

and below ground communities, carbon storage and air pollution removal (U.S. EPA, 2014b, 

Figure 5-1), that have the potential to be adverse to the public welfare. 

 Thus, in staff’s view, the evidence- and exposure/risk-based information relevant to tree 

biomass loss and the associated ecosystem services important to the public welfare support 

 

25 We note that a W126 index value of 19 ppm-hrs is estimated to result in a median RBL value of 6%. 
6-56 

consideration of a W126-based secondary standard with index values within the range of 7-17 

ppm-hrs. We consider such a range for a potential alternative cumulative seasonal W126-based 

standard, averaged over three years, based on our analysis of the small effect of year to year 

variability on the cumulative biomass loss associated with multiple years of exposure, and the 

benefits of improved stability of the W126 standard when evaluated using the 3-year average 

form. Lastly, we are mindful of the policy judgments required of the Administrator with regard 

to the public welfare significance of identified effects and the requisite level of protection, as 

well as the appropriate weight to assign the range of uncertainties inherent in the evidence and 

analyses. 

While we additionally recognize foliar injury as an important O3 effect which, depending 

on severity and spatial extent, most particularly in nationally protected areas such as Class I 

areas, may reasonably be concluded to be of public welfare significance, we take note of the 

appreciable variability in this endpoint, as summarized in chapter 5 and section 6.3 above, which 

poses challenges to giving it primary emphasis in identifying potential alternative standard 

levels. Similarly, we give less emphasis to consideration of crop yield loss in our consideration 

of potential standard levels here and in section 6.3 above, noting the median estimates of 

approximately 5% or lower for W126 index levels at and below 17 ppm-hrs. We also note the 

range of factors affecting annual crop yields, including those related to the role of management 

strategies as recognized in sections 5.3 and 6.2 above which complicate the identification of a 

degree of impact that can be considered adverse to the public welfare. 

 We further recognize the role of policy judgments by the Administrator, as described 

above, in identifying a target level of protection for the secondary O3 standard. For example, to 

the extent effects associated with cumulative multi-year exposures are judged important to the 

public welfare, more weight may be placed on such effects, as well as the role that year-to-year 

exposure variability can play in realizing the potential public welfare impacts. 

Lastly, we also conclude that, to the extent the Administrator finds it useful to consider 

the public welfare protection that might be afforded by a revised primary standard, this is 

appropriately judged by evaluating how the cumulative seasonal W126-based exposure metric is 

affected by attainment with such a revised primary standard. For example, comparison of the air 

quality conditions expected to result from a revised primary standard, with those conditions 

expressed in terms of W126 exposures, to the W126 levels concluded to provide the desired level 

of public welfare protection could inform a judgment of whether a secondary standard set 

identical to a revised primary standard would be expected to achieve the level of public welfare 

protection concluded to be requisite under the Act. In this type of evaluation, such as through 

the overlap analyses discussed in section 6.4 above, staff further concludes it is important to take 

into account associated uncertainties, including those associated with the limited monitor 
6-57 

coverage in many rural areas, including those in the west and southwest and at high elevation 

sites. 

6.7 SUMMARY OF CONCLUSIONS ON THE SECONDARY STANDARD 

Staff conclusions are informed by our consideration of the available scientific evidence as 

assessed in the ISA, the air quality/exposure/risk information in the WREA, advice from 

CASAC in this review and in prior reviews, and public comment in this review. 

Staff conclusions on policy options that are appropriate for the Administrator’s 

consideration in making decisions on the secondary standards for O3, together with supporting 

conclusions from sections 5.7 and 6.5 above, are briefly summarized below. In reaching 

conclusions on alternative standards to provide requisite protection for public welfare effects 

associated with ambient O3 exposures, staff has considered these standards in terms of the basic 

elements of the NAAQS: indicator, form, averaging time, and level. In drawing these 

conclusions, we are mindful that the Act requires secondary standards to be set so that, in the 

Administrator’s judgment, they are requisite to protect public welfare from known or anticipated 

adverse effects, such that the standards are to be neither more nor less stringent than necessary. 

Thus, the Act does not require that NAAQS be set at zero-risk or background levels, but rather at 

levels that reduce risk sufficiently to protect public welfare from adverse effects. 

(1) Staff concludes, based on the combined consideration of the body of evidence and 

the results from the quantitative exposure/risk assessment, that the available 

evidence and exposure/risk information call into question the adequacy of the 

public welfare protection provided by the current standard and that it is 

appropriate to consider revising the standard to provide greater public welfare 

protection. 

(2) With regard to indicator, staff concludes that it is appropriate to continue to use 

O3 as the indicator for a standard that is intended to address welfare effects 

associated with exposure to O3, alone or in combination with related 

photochemical oxidants. Based on the available information, staff concludes that 

there is no basis for considering an alternative indicator at this time. 

(3) With regard to averaging time and form, staff concludes that it is appropriate to 

consider a revised secondary standard in terms of the cumulative, seasonal, 

concentration-weighted form, the W126 index. With regard to definition of the 

W126 index for this purpose, staff makes the additional conclusions: 

a. It is appropriate to consider the consecutive 3-month period within the O3

season with the maximum index value as the seasonal period over which 
6-58 

to cumulate hourly O3 exposures. Staff notes that the maximum 3-month 

period generally coincides with maximum biological activity for most 

vegetation, making the 3-month duration a suitable surrogate for longer 

growing seasons. 

b. It is appropriate to cumulate daily exposures for the 12-hour period from 

8:00 am to 8:00 pm, generally representing the daylight period during the 

3-month period identified above. 

c. It is appropriate to consider a form that averages W126 index values 

across three consecutive years. Staff concludes it is appropriate to 

consider this form in conjunction with appropriate levels in order to 

provide the desired degree of public welfare protection from O3 effects 

across multiple years. 

(4) With regard to a target level of protection for a revised standard, staff concludes 

that it is appropriate to give consideration to a range of levels from 17 ppm-hrs to 

7 ppm-hrs, expressed in terms of the W126 index averaged across three 

consecutive years. 

a. To the extent the Administrator finds it useful to consider the extent of 

public welfare protection that might be afforded by a revised primary 

standard, staff concludes that public welfare protection is appropriately 

judged through the use of the cumulative seasonal W126-based metric. 

Staff additionally notes that, consideration of the support provided by the information 

available in this review will depend on public welfare policy judgments by the Administrator 

regarding the protection of public welfare. This range reflects staff judgment that a standard set 

within this range could provide an appropriate degree of public welfare protection. 

6.8 KEY UNCERTAINTIES AND AREAS FOR FUTURE RESEARCH AND DATA 

COLLECTION 

Staff believes it is important to highlight key uncertainties and recommendations for 

welfare-related research, including model development and data gathering, associated with 

secondary standards for O3. Based on items highlighted in chapter 9 of the ISA, chapters 5 and 6 

herein, and CASAC advice, we have identified the following areas for future research and data 

collection based on key uncertainties, research questions and data gaps that have been 

highlighted in this review of the secondary standard. The first research area addresses the key 

uncertainties associated with the extrapolation to plant species and environments outside of 

specific experimental or field study conditions. The second area of research pertains to the 
6-59 

assessment of the impact of O3 on other welfare effects categories such as climate, ecosystem 

components, and whole ecosystem structure and function. A third area of research would 

support the development of approaches, tools, or methodologies useful in characterizing O3

exposures in rural, remote, high elevation and/or complex terrain areas and in characterizing 

ecosystem services and their importance to the public welfare. These three areas are described 

below. 

With regard to the first research area we note that while there have been five decades of 

research regarding O3 effects on plants and much information has been compiled in previous 

reviews, a number of key uncertainties remain. For example, while national visible foliar injury 

surveys can indicate how widespread O3 effects may be within the U.S., there remain 

uncertainties associated with estimating the risk to vegetation of differing amounts of O3-induced 

visible foliar injury over the plant’s leaf area and the relationship between relative soil moisture 

and the incidence and severity of foliar injury in sensitive species, as well as the extent to which 

different degrees of visible foliar injury can impact ecosystem services (e.g., tourism). Research 

to better characterize the relationship between O3, soil moisture and foliar injury and to 

determine if there is an injury threshold or quantifiable relationship between these factors could 

help inform policy. Additionally, research to understand the connection between O3-related 

foliar injury and other physiological effects and ecosystem services could also be useful. We 

further note that while this review relied on the robust E-R functions that are available for 11 

tree and 10 crop species, there are tens of thousands of plant species in the U.S. (USDA, NRCS, 

2014),26 66 of which have also been identified as O3 sensitive on National Park Service and US 

Fish and Wildlife Service lands27. Research on additional tree as well as non-tree species that 

would support the development of robust E-R functions would improve our understanding of the 

full range of response of plant species to O3 and our understanding of the overall risk to 

vegetation. For example, studies using large numbers of native plant species across regions 

where those species are indigenous, might be expected to reduce uncertainties associated with 

extrapolating plant response for a given level of O3 using composite response functions across 

differing regions and climates. Studies focused on fruits and vegetables might assist in reducing 

uncertainties associated with O3 effects on agriculture. Particular focus is suggested on 

organically grown vegetables that may receive less intensive management than conventionally 

grown crops. Recent studies indicate that watermelons may be particularly sensitive to O3

 

26 USDA, NRCS. 2014. The PLANTS Database (http://plants.usda.gov, 3 January 2014). National Plant 

Data Team, Greensboro, NC 27401-4901 USA. 

27See http://www2.nature.nps.gov/air/Pubs/pdf/flag/NPSozonesensppFLAG06.pdf 

 
6-60 

exposure (U.S. EPA, 2013, section 9.4.4.1) and older studies indicate grapes, honeydew melon, 

lemons and oranges may also be O3 sensitive (Abt Associates Inc., 1995). 

Some new information has emerged linking effects on tree seedlings with larger trees and 

similarities in results between exposure techniques (U.S. EPA 2013, section 9.6). Uncertainties 

remain in this area as well as uncertainties in extrapolating from O3 effects on juvenile to mature 

trees and from trees grown in the open versus those in a closed forest canopy in a competitive 

environment. The relationship between nocturnal exposures and plant uptake and response is 

also an important subject for further research. 

With respect to the second research area pertaining to the impact of O3 on other welfare 

effects categories such as climate, ecosystem components, and whole ecosystem structure and 

function, uncertainties that remain in extrapolating individual plant response spatially or to 

higher levels of biological organization, including ecosystems, could be informed by research 

that explores and better quantifies the nature of the relationship between O3, plant response and 

multiple biotic and abiotic stressors, including those associated with the ecosystem services that 

would be affected (e.g., hydrology, productivity, carbon sequestration). Because these 

uncertainties are multiple and significant due to the complex interactions involved, new research 

will likely 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. 

Uncertainties associated with projections of the effects of O3 on the ecosystem processes 

of water, carbon, and nutrient cycling, particularly at the stand and community levels might be 

addressed through research on the effects on below ground ecosystem processes in response to 

O3 exposure alone and in combination with other stressors. These below-ground processes 

include interactions of roots with the soil or microorganisms, effects of O3 on structural or 

functional components of soil food webs and potential impacts on plant species diversity, 

changes in the water use of sensitive trees, and if the sensitive tree species is dominant, potential 

changes to the hydrologic cycle at the watershed and landscape level. Research on competitive 

interactions under different O3 exposures might improve our understanding of how O3 may affect 

biodiversity or genetic diversity. Such research could be strengthened by modern molecular 

methods to quantify impacts on diversity. More tools and research would improve our 

understanding of relationships between O3 exposure and stressors such as insect infestations, 

plant diseases, drought and potential stressors from climate change. It is also important to 

understand how such interactions may affect ecosystem services such as CO2 sequestration; food 

and fiber production; wildlife habitat and water resources. 

With respect to the research areas related to the development of approaches, tools, or 

methodologies useful in characterizing O3 exposures and the relationship between O3-induced 
6-61 

effects and associated ecosystem services and public welfare in a policy context, we note that 

one of the most important uncertainties in this review is the characterization of air quality in rural 

areas where there is limited monitoring. More comprehensive monitoring in these areas would 

reduce uncertainties associated with O3 exposures in many rural areas. Areas of particular 

uncertainty include protected natural areas in the western U.S, including those at high elevation, 

as well as those downwind of recently expanded oil and gas development areas. Uncertainties 

associated with quantifying exposure in areas with and without monitors might be addressed 

through additional work on interpolation methods and air quality models that are tailored to 

estimating cumulative seasonal exposures, as well as improved model capabilities that use more 

refined spatial grids and are better able to handle O3 movement in complex terrain. 

Uncertainties related to characterizing the potential public welfare significance of O3-

induced effects and impacts to associated ecosystem services could also be informed by research, 

such as research intended to clarify the relationship between O3 exposure and fire risk and O3

exposure and forest susceptibility to bark beetle infestation. Research relating known O3

ecological effects such as reproductive effects to effects on production of non-timber forest 

products and research to characterize public preferences including valuation related to non-use 

and recreation for foliar injury could also help inform consideration of the public welfare 

significance of these effects. 
6-62 

6.9 REFERENCES 

Abt Associates, Inc. (1995). Ozone NAAQS benefits analysis: California crops. Report to U.S. EPA, July. 

Federal Register 1996. National Ambient Air Quality Standards for Ozone; Proposed Rule. 40 CFR 50; Federal 

Register 61: 65716 

Federal Register 1997. National Ambient Air Quality Standards for Ozone; Final Rule. 40 CFR 50; Federal Register 

62:38856 

Federal Register 2007. National Ambient Air Quality Standards for Ozone; Proposed Rule. 40 CFR 50; Federal 

Register 72: 37818 

Federal Register 2008. National Ambient Air Quality Standards for Ozone; Final Rule. 40 CFR 50 and 58; Federal 

Register 73:16436 

Federal Register 2010. National Ambient Air Quality Standards for Ozone; Proposed Rule. 40 CFR 50 and 58; 

Federal Register 75 FR: 2938 

Federal Register 2012. National Ambient Air Quality Standards for Oxides of Nitrogen and Sulfur; Final Rule. 40 

CFR 50; Federal Register 77 FR 20218 

Frey, C., (2014a) CASAC Review of the EPA’s Second Draft Policy Assessment for the Review of the Ozone 

National Ambient Air Quality Standards. EPA-CASAC-14-004. June 26, 2014. 

Frey, C. (2014b) CASAC Review of the EPA’s Welfare Risk and Exposure Assessment for Ozone (Second 

External Review Draft). EPA-CASAC-14-003. June 18, 2014. 

Gregg, JW; Jones, CG; Dawson, TE. (2003). Urbanization effects on tree growth in the vicinity of New York City 

[Letter]. Nature 424:183-187. http://dx.doi.org/10.1038/nature01728 

Heck, WW; Cowling, EB. (1997).The need for a long term cumulative secondary ozone standard - An ecological 

perspective. EM January: 23-33. 

Henderson, R. (2006) Letter from CASAC Chairman Rogene Henderson to EPA Administrator Stephen Johnson. 

October 24, 2006, EPA-CASAC-07-001. 

Henderson, R. (2007) Letter from CASAC Chairman Rogene Henderson to EPA Administrator Stephen Johnson. 

March 26, 2007, EPA-CASAC-07-002. 

Henderson, R. (2008) Letter from CASAC Chairman Rogene Henderson to EPA Administrator Stephen Johnson. 

April 7,2008, EPA-CASAC-08-009. 

Lefohn, AS; Laurence, JA; Kohut, RJ. (1988). A comparison of indices that describe the relationship between 

exposure to ozone and reduction in the yield of agricultural crops. Atmos Environ 22: 1229-1240. 

http://dx.doi.org/10.1016/0004-6981(88)90353-8 

Lefohn, AS; Runeckles, VC. (1987). Establishing standards to protect vegetation - ozone exposure/dose 

considerations. Atmos Environ 21: 561-568. http://dx.doi.org/10.1016/0004-6981(87)90038-2 

Samet, J.M. (2010) Clean Air Scientific Advisory Committee (CASAC) Review of EPA’s Proposed Ozone National 

Ambient Air Quality Standard. EPA-CASAC-10-007. January 19, 2010. Available online at: 

http://yosemite.epa.gov/sab/sabproduct.nsf/264cb1227d55e02c85257402007446a4/610BB57CFAC8A41C

852576CF007076BD/$File/EPA-CASAC-10-007-unsigned.pdf 
6-63 

USDA, NRCS. 2014. The PLANTS Database (http://plants.usda.gov, 3 January 2014). National Plant Data Team, 

Greensboro, NC 27401-4901 USA 

U.S. EPA (1996a). Air quality criteria for ozone and related photochemical oxidants [EPA Report]. (EPA/600/P93/004AF). U.S. Environmental Protection Agency, Research Triangle Park, NC. 

U.S. EPA (1996b). Review of national ambient air quality standards for ozone: Assessment of scientific and 

technical information: OAQPS staff paper [EPA Report]. (EPA/452/R-96/007). Research Triangle Park, 

NC. http://www.ntis.gov/search/product.aspx?ABBR=PB96203435 

U.S. EPA (2006). Air Quality Criteria for Ozone and Related Photochemical Oxidants (2006 Final). U.S. 

Environmental Protection Agency, Washington, DC. EPA/600/R-05/004aF-cF. March 2006. Available at: 

http://www.epa.gov/ttn/naaqs/standards/ozone/s_o3_cr_cd.html 

U.S. EPA (2007). Review of the national ambient air quality standards for ozone: Policy assessment of scientific and 

technical information: OAQPS staff paper [EPA Report]. (EPA/452/R-07/003). Research Triangle Park, 

NC. http://www.epa.gov/ttn/naaqs/standards/ozone/data/2007_01_ozone_staff_paper 

U.S. EPA (2013). Integrated Science Assessment of Ozone and Related Photochemical Oxidants (Final). U.S. 

Environmental Protection Agency, Washington, DC. EPA/600/R-10/076F 

U.S. EPA (2014a). Policy Assessment for the Review of National Ambient Air Quality Standards for Ozone, 

Second External Review Draft. Office of Air Quality Planning and Standards, Research Triangle Park, NC, 

27711. EPA–452/P–14–002 

U.S. EPA (2014b). Welfare Risk and Exposure Assessment for Ozone, Second External Review Draft. Office of 

Air Quality Planning and Standards, Research Triangle Park, NC, 27711. EPA-452/P-14-003a 

U.S. Forest Service, National Park Service, and U.S. Fish and Wildlife Service. 2010. Federal land managers’ air 

quality related values work group (FLAG): phase I report—revised (2010). Natural Resource Report 

NPS/NRPC/NRR—2010/232. National Park Service, Denver, Colorado. 

Wolff, G.T. (1996) Letter to EPA Administrator Carol Browner, RE: Closure by the Clean Air Scientific Advisory 

Committee (CASAC) on the Secondary Standard Portion of the Staff Paper for Ozone. EPA-SAB-CASACLTR-96-002, April 4, 1996. 
APPENDICES 

Appendix 2A. Supplemental Air Quality Modeling Analyses of Background O3........................2A-1 

Appendix 2B. Monitoring Data Analysis of Relationships Between Current Standard and W126 

Metric.................................................................................................................2B-1 

Appendix 2C. Inter-annual Variability in W126 Index Values: Comparing Annual and 3-Year Average 

Metrics (2008-2010)...............................................................................................2C-1 

Appendix 3A. Recent Studies of Respiratory-Related Emergency Department Visits and Hospital 

Admissions..........................................................................................................3A-1 

Appendix 3B. Ambient O3 Concentrations in Locations of Health Studies................................3B-1 

Appendix 5A. O3-Sensitive Plant Species Used by Some Tribes...........................................5A-1 

Appendix 5B. Class I Areas Below Current Standard And Above 15 ppm-hrs...........................5B-1 

Appendix 5C. Expanded Evaluation of Relative Biomass and Yield Loss.................................5C-1 
2A-1 

APPENDIX 2A 

SUPPLEMENTAL AIR QUALITY MODELING ANALYSES OF 

BACKGROUND OZONE 

Table of Contents

List of Figures  ............................................................................................................................ 2

List of Tables  .............................................................................................................................. 4   

1.   Introduction  ..................................................................................................................   5

2.   Description of modeling methodologies  .................................................................. 6

a.  2007 GEOS‐Chem/CMAQ zero‐out modeling ........................................................... 7

b.  2007 GEOS‐Chem/CAMx source apportionment modeling .............................. 12

3. Estimates of seasonal‐average background ozone levels  ....................................... 15

4. Distributions of background ozone levels  .................................................................. 22

5. Contribution of various processes and sources to total background ozone  ......    30

6. Estimates of the fractional background contribution to total ozone in 12 specific  

areas   .............................................................................................................................. 37

7. Background ozone and W126  ..................................................................................... 39

8. Summary  ........................................................................................................................    39

9.   References  .....................................................................................................................    42

   
2A-2 

List of Figures

Figure 1a.  Modeling domain used in 2007 CMAQ and CAMx modeling  ........................................... 10

Figure 1b.  Density scatterplot comparing CMAQ base daily peak 8‐hour ozone predictions  

against observed 8‐hour ozone peaks paired in space and time for all sites during April‐ 

October 2007 ................................................................................................................................................ 10

Figure 1c.  Bias in seasonal mean (April‐October) maximum daily 8‐hour ozone predictions in  

the 2007 CMAQ base simulation  ............................................................................................................... 11

Figure 1d.  Relationship between CMAQ estimations of MDA8 natural background ozone and  

daily model biases .........................................................................................................................................11

  

Figure 2a.  Density scatterplot comparing CAMx base daily peak 8‐hour ozone predictions  

against observed 8‐hour ozone peaks paired in space and time for all sites during April‐October  

2007 ................................................................................................................................................................ 14  

Figure 2b.  Bias in seasonal mean (April‐October) maximum daily 8‐hour ozone predictions in  

the 2007 CAMx base simulation ................................................................................................................. 14

Figure 3a.  April‐October average MDA8 ozone (ppb) at monitoring locations across the U.S. as

estimated by a 2007 CMAQ base simulation ............................................................................................18

Figure 3b.  April‐October average natural background MDA8 ozone (ppb) at monitoring  

locations across the U.S. as estimated by a 2007 CMAQ zero out simulation .................................... 18

Figure 3c.  April‐October average North American background MDA8 ozone (ppb) at monitoring

locations across the U.S. as estimated by a 2007 CMAQ zero out simulation .....................................19

Figure 3d.  April‐October average United States background MDA8 ozone (ppb) at monitoring  

locations across the U.S. as estimated by a 2007 CMAQ zero out simulation .....................................19

Figure 4a.  Ratio of natural background to total April‐October average MDA8 ozone at  

monitoring locations across the U.S. as estimated based on 2007 CMAQ simulations ..................... 20

Figure 4b.  Ratio of N. American background to total April‐October average MDA8 ozone at  

monitoring locations across the U.S. as estimated based on 2007 CMAQ simulations ..................... 20

Figure 4c.  Ratio of U.S.  background to total April‐October average MDA8 ozone at  

monitoring locations across the U.S. as estimated based on 2007 CMAQ simulations ..................... 21

Figure 4d.  Ratio of sources other than U.S. anthropogenic emissions to total April‐October  

average MDA8 ozone at monitoring locations across the U.S. as estimated by a 2007 CAMx  

source apportionment simulation ..............................................................................................................21  

Figure 5a.  Distribution of natural background MDA8 ozone (ppb) at monitoring locations  

across the U.S. (Apr‐Oct), binned by base modeled site‐day MDA8, as estimated by 2007  
2A-3 

CMAQ simulations ..................................................................................................................................... 25

Figure 5b.  Distribution of N. American background MDA8 ozone (ppb) at monitoring  

locations across the U.S. (Apr‐Oct), binned by base modeled site‐day MDA8, as estimated by  

2007 CMAQ simulations ............................................................................................................................ 25

Figure 5c.  Distribution of U.S. background MDA8 ozone (ppb) at monitoring locations across  

the U.S. (Apr‐Oct), binned by base modeled site‐day MDA8, as estimated by 2007 CMAQ  

simulations .................................................................................................................................................. 26

Figure 5d.  Distribution of  MDA8 ozone contributions from non‐U.S. manmade sources (ppb)  

at monitoring locations across the U.S. (Apr‐Oct), binned by base modeled site‐day MDA8, as  

estimated by 2007 CAMx simulations ..................................................................................................... 26

Figure 6a.  Distribution of natural background MDA8 ozone fractions at monitoring locations  

across the U.S. (Apr‐Oct), binned by base modeled site‐day MDA8, as estimated by 2007  

CMAQ simulations ...................................................................................................................................... 27

Figure 6b.  Distribution of N. American background MDA8 ozone fractions at monitoring  

locations across the U.S. (Apr‐Oct), binned by base modeled site‐day MDA8, as estimated by  

2007 CMAQ simulations ............................................................................................................................. 27

Figure 6c.  Distribution of U.S. background MDA8 ozone fractions at monitoring locations  

across the U.S. (Apr‐Oct), binned by base modeled site‐day MDA8, as estimated by 2007  

CMAQ simulations ....................................................................................................................................... 28

Figure 6d.  Distribution of MDA8 ozone fractions from non‐U.S. anthropogenic sources at  

monitoring locations across the U.S. (Apr‐Oct), binned by base modeled site‐day MDA8, as  

estimated by the 2007 CAMx simulation .................................................................................................. 28

Figure 7.  April‐October 95th percentile United States background MDA8 ozone (ppb) at  

monitoring locations across the U.S. as estimated by a 2007 CMAQ base simulation ..................... 29

Figure 8a.  Difference in April‐October average MDA8 ozone (ppb) at monitoring locations  

across the U.S. between the USB scenario and the NAB scenario  ...................................................... 32

Figure 8b.  Difference in April‐October average MDA8 ozone (ppb) at monitoring locations  

across the U.S. between the NAB scenario and the NB scenario .......................................................... 32  

Figure 9a.  Percentage of April‐October average MDA8 ozone that is apportioned to  

boundary conditions  as estimated at monitoring locations by a 2007 CAMx simulation ................. 33

Figure 9b.  Percentage of April‐October average MDA8 ozone that is apportioned to U.S.

anthropogenic sources  as estimated at monitoring locations by a 2007 CAMx simulation ............ 33

Figure 9c.  Percentage of April‐October average MDA8 ozone that is apportioned to purely  

biogenic emissions  as estimated at monitoring locations by a 2007 CAMx simulation .................... 34
2A-4 

Figure 9d.  Percentage of April‐October average MDA8 ozone that is apportioned to  

climatological fire emissions  as estimated at monitoring locations by a 2007 CAMx simulation ... 34

Figure 9e.  Percentage of April‐October average MDA8 ozone that is apportioned to  

anthropogenic emissions from in‐domain Canadian and Mexican sources as estimated at  

monitoring locations by a 2007 CAMx simulation ....................................................................................35

Figure 9f.  Percentage of April‐October average MDA8 ozone that is apportioned to Category 3  

marine vessel emissions beyond U.S. territorial waters as estimated at monitoring locations by  

a 2007 CAMx simulation .............................................................................................................................. 35

Figure 9g.  Percentage of April‐October average MDA8 ozone that is apportioned to Gulf of  

Mexico point sources as estimated at monitoring locations by a 2007 CAMx simulation ............... 36

List of Tables

Table 1a.  April‐October average MDA8 ozone, average MDA8 ozone from sources other than  

U.S. manmade emissions, and the fractional contribution of these background sources in the  

12 REA urban study areas, as estimated by a 2007 CAMx simulation................................................. 37

Table 1b.  Average MDA8 ozone, average MDA8 ozone from sources other than U.S.  

manmade emissions, and the fractional contribution of these background sources in the 12 REA  

areas, as estimated by a 2007 CAMx simulation using site‐days in which base MDA8 ozone  

exceeded 60 ppb .......................................................................................................................................... 38

Table 1c.  Fractional contribution of non‐U.S. manmade emissions sources in the 12 REA urban  

study areas, as estimated by a 2007 CAMx simulation using means and medians of daily MDA8  

fractions ......................................................................................................................................................... 38

Table 1d.  April‐October average MDA8 ozone, average MDA8 ozone from USB, and the  

fractional contribution of these background sources in the 12 REA urban study areas, as  

estimated by two separate 2007 CMAQ simulations .............................................................................. 38

   
2A-5 

1. Introduction

One of the aspects of ozone that is unusual relative to the other pollutants with National

Ambient Air Quality Standards (NAAQS) is that, periodically, in some locations, an appreciable fraction of

the observed ozone results from sources or processes other than local and regional anthropogenic

emissions of ozone precursors (Fiore et al., 2002).  Any ozone formed by processes other than the

chemical conversion of local or regional ozone precursor emissions, such as nitrogen oxides (NOx) or

volatile organic emissions (VOC), is generically referred to as “background” ozone.  As part of this review

of the ozone NAAQS, EPA completed an extensive review of the known aspects of background ozone

and summarized the findings in the Integrated Science Assessment (ISA) in March 2013 (USEPA, 2013).  

The purpose of this appendix is to present the results from supplemental air quality modeling analyses

related to background ozone that were completed by EPA subsequent to the ISA.  While these updated

analyses use a recent base year (2007) and consider an alternative modeling methodology which can

better account for non‐linear ozone chemistry in some conditions, the results are largely consistent with

previous determinations about the magnitude of background ozone contributions across the U.S.  

Away from the surface, ozone can have an atmospheric lifetime on the order of weeks.  As a

result, background ozone can be transported long distances at heights above the boundary layer and,

when meteorological conditions are favorable, be available to mix down to the surface and add to the

total ozone loading from non‐background sources.  Generically, background ozone can originate from

natural sources of ozone and ozone precursors, as well as from far upwind manmade emissions of ozone

precursors.   Natural sources of ozone precursor emissions such as wildfires, lightning, and vegetation

can lead to ozone formation by chemical reactions with other natural sources1

.  Another important

natural component of background is ozone that is naturally formed in the stratosphere through

interactions of UV light with atomic oxygen (O2).  Stratospheric ozone can periodically mix down to the

surface at high concentrations, especially at higher altitude locations.  The manmade portion of the

background includes any ozone formed due to anthropogenic sources of ozone precursors emitted far

away from the local area (e.g., international emissions).  Finally, both biogenic and international

anthropogenic emissions of methane, which can be chemically converted to ozone over relatively long

time scales, can also contribute to global background ozone levels.

The precise definition of background ozone can vary depending upon context, but it generally

refers to ozone that is formed by sources or processes that cannot be influenced by actions within the

jurisdiction of concern.  In the first draft policy assessment document (EPA, 2012), EPA presented three

specific definitions of background ozone: natural background, North American background, and U.S.

background.  Natural background (NB) was the narrowest definition of background and it was defined as

the ozone that would exist in the absence of any manmade ozone precursor emissions.  The other two

previously‐established definitions of background presume that the U.S. has little influence over

anthropogenic emissions outside our continental or domestic borders.  North American background

(NAB) is defined as that ozone that would exist in the absence of any manmade ozone precursor

                                                            

1

  Ozone formed through reactions between natural emissions and local anthropogenic emissions (e.g., biogenic

VOC with man‐made NOx) is generally not considered to be background ozone.
2A-6 

emissions inside of North America.  U.S. background (USB) is defined as that ozone that would exist in

the absence of any manmade emissions inside the United States.  It is important to note that each of

these three definitions of background ozone requires photochemical modeling simulations to estimate

what the residual ozone concentrations would be were the various anthropogenic emissions to be

removed.   

As noted in the first draft policy assessment, EPA has revised several aspects of our

methodology for estimating the change in health risk and exposure that would result from a revision to

the ozone NAAQS.  First, risk estimates are now based on total ozone concentrations as opposed

previous reviews which only considered risk above background levels.  Second, EPA is now using air

quality models to estimate the spatial patterns of ozone that would result from attaining various levels

of the NAAQS, as opposed to simplistic rollback techniques that required the estimation of a background

ozone “floor” beyond which the rollback would not take place.  Both of these revisions have had the

indirect effect of obviating the need for estimating background ozone levels as part of the ozone risk

and exposure assessment (REA).  Regardless, EPA expects that a well‐founded understanding of the

fractional contribution of background sources and processes to surface ozone levels will be valuable

towards informing policy decisions about the NAAQS.  Section 2 of this document will describe the

supplemental air quality modeling simulations that have recently been completed by EPA to bolster our

understanding of background ozone.  Section 3 will present the results from the updated analyses and

provide estimates of average background ozone levels, and how they can vary in time and space across

the U.S.  Based on the same modeling, Section 4 will consider the entire spectrum of variable

background ozone levels with special emphasis on areas and times in which background can approach or

exceed the level of the NAAQS.  Section 5 will utilize the supplemental air quality modeling estimates to

determine the relative importance of specific components of background ozone.  Section 6 will present

estimates of the overall fraction of ozone that is estimated to result from background sources or

processes in each of the 12 urban case study areas in the epidemiology study based analyses in Chapter

7 of the Risk and Exposure Assessment (REA) (EPA, 2014) based on the updated modeling.  Finally,

Section 7 will conclude with a limited analysis of how background ozone levels impact longer‐term

ozone metrics that may be important from a welfare perspective (i.e., W126).

2. Description of modeling methodologies

As noted above, air quality models are typically used to estimate background ozone as it is quite

difficult to measure directly.  Without special monitoring, it is impossible to determine how much of the

ozone measured by a monitor originated from sources that are considered background.  Even the most

remote monitors within the U.S. can periodically be affected by U.S. anthropogenic emissions.  Previous

modeling studies have estimated what background levels would be in the absence of certain sets of

emissions by simply comparing the ozone differences between a base model simulation and a control

simulation in which emissions were removed.  This basic approach is often referred to as “zero out”

modeling or “emissions perturbation” modeling.  Examples of zero out modeling include the three major

studies summarized in the ISA (Zhang et al., 2011; Emery et al., 2012, Lin et al., 2013).  It is important to

note that the specific concepts of NB, NAB, and USB are all explicitly tied to zero‐out modeling, as those

definitions are based on estimating what remains in the absence of specific sets of man‐made emissions.  
2A-7 

EPA has conducted and will describe updated air quality modeling for a 2007 base year that employs a

regional air quality model nested within a coarser‐scale global chemical transport model to estimate NB,

NAB, and USB levels when the respective manmade emissions are zeroed.  This modeling is described in

detail in section 2a.

While the zero‐out approach has traditionally been used to estimate background ozone levels,

the methodology has some acknowledged limitations.  First, from a policy perspective, the purely

hypothetical and ultimately unrealizable zero manmade emissions scenarios have limited application in

this regard.  Secondly, the assumption that background ozone is what is left after specific emissions have

been removed within the model simulation can be misleading in locations where ozone chemistry is

highly non‐linear.  Depending upon the local composition of ozone precursors, NOx emissions

reductions can either increase or decrease ozone levels in the immediate vicinity of those reductions.  

For those specific urban areas in which NOx titration of ozone can be significant, zero‐out modeling can

result in inflated estimates of background ozone when these NOx emissions are completely and

unrealistically removed.  Paradoxically, in certain times and locations in a zero‐out scenario there can be

more background ozone than actual ozone within the model (EPA, 2014).    

A separate modeling technique attempts to circumvent these limitations by apportioning the

total ozone within the model to its contributing source terms.  This basic approach is referred to as

“source apportionment” modeling.  While source apportionment modeling has not been previously used

in the context of estimating background ozone levels as part of an ozone NAAQS review, it has

frequently been used in other regulatory settings to estimate the “contribution” to ozone of certain sets

of emissions (EPA 2005, EPA 2011).  The source apportionment technique provides a means of

estimating the contributions of user‐identified source categories to ozone formation in a single model

simulation.  This is achieved by using multiple tracer species to track the fate of ozone precursor

emissions (VOC and NOx) and the ozone formation caused by these emissions.  The methodology is

designed so that all ozone and precursor concentrations are attributed to the selected source categories

at all times without perturbing the inherent chemistry.  The zero out modeling attempts to determine

what ozone be in the absence of background sources.  The source apportionment modeling attempts to

determine how much of the modeled ozone has resulted from background sources.  EPA has conducted

and will describe new source apportionment modeling that employs a regional air quality model nested

within a coarser‐scale global chemical transport model to assess the contributions of boundary

conditions and other potential background sources (e.g., wildfires, biogenic emissions, and

Canadian/Mexican emissions).  This modeling is described in detail in section 2b.

a.  2007 GEOS‐Chem/CMAQ zero‐out modeling:   

In order to provide estimates of the overall fraction of ozone that is estimated to result from

background sources in each of the 12 REA urban study areas, EPA conducted new modeling that utilized

the same model base year (2007) as was used in the ozone modeling that inform the risk and exposure

analyses (EPA, 2014, Appendix 4b).  The EPA modeling used a model configuration similar to that of

Emery (2012), in that it nested a regional‐scale (12 km) air quality model inside a global air quality model
2A-8 

simulation with a much coarser horizontal grid resolution (2.0 by 2.5 degrees).  Figure 1a shows a map of

the model domain.

The global scale simulation utilized the GEOS‐Chem model, version v8‐03‐02, except for the

chemistry package which was from version v8‐02‐01.  The emissions estimates used in the 2007 base

year modeling were aggregated from a variety of sources, starting with the global Emissions Database

for Global Atmospheric Research (EDGAR) emission inventory.  These initial estimates were then

improved by utilizing various area‐specific inventories, such as the 2005 National Emissions Inventory

(NEI) for the U.S. portions of the domain, and available inventories for Asia, Canada, Europe, and

Mexico.  In addition to the anthropogenic estimates, emissions were specified for a variety of

background sources including: lightning NO, soil NOx, wildfires, and biogenic VOC emissions.  The

wildfire data is from the Global Fire Emissions Database (GFED).  The biogenic VOC estimates were

simulated by the Model of Emissions of Gases and Aerosols from Nature (MEGAN) version 2.1.  The

meteorological data is based on the Goddard Earth Observing System Model, Version 5 (GEOS‐5)

analysis fields.  More information on the global simulation is available within Henderson et al. (2013).  

This reference also provides a broad evaluation of the ability of this specific GEOS‐Chem configuration to

provide accurate lateral boundary conditions of ozone to finer‐scale regional simulations.  Using satellite

retrievals from the Tropospheric Emissions Spectrometer (TES), Henderson et al. (2013) concluded that

the GEOS‐Chem ozone prediction biases and errors are generally within TES uncertainty estimates.  For

instance, for the ozone season month of August, model predictions are within plus or minus 20 percent

of the satellite estimates between nearly 80 percent of the time, with slightly better performance along

the southern boundary.

The lateral boundary conditions from the global model were then used as inputs for a 12 km

horizontal resolution, CMAQ version 4.7.1, model simulation.  Four scenarios were modeled: 1) a 2007

base case simulation which was the basis of the air quality modeling performed for the 2nd draft ozone

REA and is described in more detail in Appendix 4b of EPA (2014), 2) a natural background run with

anthropogenic ozone precursor emissions2 removed in both the global and regional models, 3) a North

American background run with anthropogenic ozone precursor emissions removed across North

America (global and regional model simulations), and 4) a U.S. background run with anthropogenic

ozone precursor emissions were removed over the U.S (global and regional model simulations).  

Detailed analyses of EPA’s 2007 zero out modeling results are provided in sections 3 through 6 of this

appendix.  

An operational model performance evaluation was completed for surface ozone in the 2007

base simulation as described separately (EPA 2014, Appendix 4b).   For the purposes of this analysis, EPA

assessed the model ability to reproduce measured daily maximum 8‐hour (MDA8) ozone values and

seasonal mean MDA8 ozone concentrations for the period April to October 2007.  As noted earlier, the

base year modeling in this analysis used climatological monthly‐average wildfire emissions which are not

                                                            

2

  In the global model all ozone precursor species were removed (i.e., VOC, NOx, CO), except for methane which

was reset to pre‐industrial levels to reflect natural contributions.  In the regional modeling, the methane levels

were left unchanged.  
2A-9 

intended to capture discrete events from specific fires that occurred in 2007, so perfect correlation

between observations and model predictions should not be expected.  Figure 1b provides a density

scatterplot of the observed and predicted daily 8‐hour ozone peaks paired in space and time for the

2007 CMAQ base.  As can be seen, the majority of pairs line up along the 1:1 line.  There is a tendency

for the model to overestimate site‐days with low 8‐hour ozone peaks, and underestimate the site‐days

with higher peak ozone values.  Modeled 8‐hour ozone peak concentrations exhibited relatively small

bias and error compared to the observations.  The average bias in MDA8 ozone estimates was 3.5 ppb.  

Figure 1c depicts the spatial bias patterns in MDA8 ozone at all sites that measured valid ozone data for

at least 100 days during the April‐October period.  CMAQ overestimations are greatest along the Gulf

Coast region, along the Atlantic coastline, and over the central U.S.  The majority of underestimated

seasonal mean MDA8 occurs within southern California.  The model performance for the 2007 base

simulation is equivalent or better than typical state‐of‐the‐science photochemical model performance

recently reported in the literature (Simon et al, 2012).

Certainly some remote monitoring locations are more affected by background sources than

other locations in the network.  However, this and numerous other analyses have shown that even the

most remote ozone monitoring locations in the U.S. are periodically affected by U.S. manmade

emissions.  In this analysis we carefully assess model performance to ensure that model error does not

influence the characterization of background ozone.  As noted in the recent ISA (EPA, 2013), there is

greater confidence in the ability of the model to predict mean contributions from background sources

rather than individual events.  Beyond the statistical analyses summarized in the previous paragraph and

in appendix 4b of the 2nd draft ozone REA (EPA, 2014), it is valuable to attempt to diagnose the model

ability to account for background ozone within the simulation.  EPA assessed whether any correlation

existed between daily model biases and daily background ozone estimates.  Figure 1d shows that at

high‐elevation sites (i.e., sites more than 1km above sea level) the highest estimates of natural

background ozone tend occur on days with greatest overestimation.  Conversely, the site‐days with the

lowest natural background estimates tend to occur when the model underestimates the observed daily

peaks at these sites.  This relationship between background estimates and simulation bias appears to be

constrained to the mountainous portion of the Western U.S.  Figure 1d also shows that estimates of

natural background ozone greater than 60 ppb are associated with large over‐predictions.  However,

based on the relatively low model bias and the general lack of correlation between daily bias values and

background estimates, EPA believes that these model estimates can be used to help characterize

background ozone levels over the U.S.   Although the highest background estimates should be

considered with caution.  
2A-10 

  

Figure 1a.  Modeling domain used in 2007 CMAQ and CAMx modeling.

Figure 1b.  Density scatterplot comparing CMAQ base daily peak 8‐hour ozone predictions against

observed 8‐hour ozone peaks paired in space and time for all sites during April‐October 2007.   
2A-11 

Figure 1c.  Bias in seasonal mean (April‐October) maximum daily 8‐hour ozone predictions in the 2007

CMAQ base simulation.

Figure 1d.  Relationship between CMAQ estimations of MDA8 natural background ozone and daily

model biases.  
2A-12 

b.  2007 GEOS‐Chem/CAMx source apportionment modeling:   

The same global modeling described above was used to assign lateral boundary conditions to

the regional‐scale (12 km) CAMx v5.0 source apportionment simulations.  Wherever possible, the

emissions and meteorological inputs in the CAMx modeling were chosen to mimic the 2007 base CMAQ

simulation described earlier.  Figure 1a shows a map of the model domain.

As with the CMAQ base case, a limited operational model performance evaluation was also

completed for surface ozone in the 2007 base simulation.  For the purposes of this analysis, EPA

assessed the model ability to reproduce measured daily maximum 8‐hour (MDA8) ozone values and

seasonal mean MDA8 ozone concentrations for the period April to October 2007.  Figure 2a provides a

density scatterplot of the observed and predicted daily 8‐hour ozone peaks paired in space and time for

the 2007 CAMx base simulation.  As can be seen, the majority of pairs line up along the 1:1 line.  Again,

there is a tendency for the model to overestimate site‐days with low 8‐hour ozone peaks and

underestimate the site‐days with higher peak ozone values.  Modeled 8‐hour ozone peak concentrations

exhibited relatively small bias and error compared to the observations.  The average bias in MDA8 ozone

estimates was 3.5 ppb.  Figure 2b depicts the spatial bias patterns in MDA8 ozone at all sites that

measured valid ozone data for at least 100 days during the April‐October period.  CAMx overestimations

are greatest along the Gulf Coast region, along the Atlantic and Pacific coastlines, and within the

southeastern U.S.  The majority of underestimated seasonal mean MDA8 occurs in California away from

the coastline.   

The apportionment tools in CAMx utilized here to estimate the contribution of background

sources are well‐established and have previously been peer‐reviewed (UNC, 2009).  EPA used the

Anthropogenic Precursor Culpability Assessment (APCA) tool in this analysis.  The APCA tool attributes

ozone production to manmade sources whenever ozone is determined to result from a combination of

anthropogenic and biogenic emissions (Environ, 2011).  The APCA methodology defines natural ozone as

the production resulting from the interaction of biogenic VOC with biogenic NOx emissions.  Eleven

separate source categories were tracked in the source apportionment analysis, including five boundary

condition terms and six in‐domain sectors:

 Boundary condition terms:

o Northern edge

o Eastern edge

o Southern edge

o Western edge

o Top boundary

 In‐domain sectors:

o U.S. anthropogenic emissions

o Point sources located within the Gulf of Mexico

o Category 3 marine vessels outside State boundaries

o Climatologically‐averaged wildfire emissions
2A-13 

o Biogenic emissions

o Canada/Mexico emissions (only those sources within the domain)

It should be noted that the source apportionment modeling conducted here does not allow for

replication of natural background because of the construct of boundary conditions.  The boundary

conditions for our applications can include ozone and/or ozone precursors that were originally

generated by natural sources, as well as ozone produced from far upstream anthropogenic emissions

(e.g., Asia).  It is not possible to disentangle these two terms.  Instead, the source apportionment

modeling is primarily used to help estimate background into the U.S., which is assumed to be the

contributions from nine of the modeled sectors; that is, everything except U.S. anthropogenic emissions

and point sources located within the Gulf of Mexico.

Figure 2a.  Density scatterplot comparing CAMx base daily peak 8‐hour ozone predictions against

observed 8‐hour ozone peaks paired in space and time for all sites during April‐October 2007.  
2A-14 

Figure 2b.  Bias in seasonal mean (April‐October) maximum daily 8‐hour ozone predictions in the 2007

CAMx base simulation.

3. Estimates of seasonal‐average background ozone levels

This section of the appendix provides estimates of seasonal average background ozone levels

over the U.S.  As noted in the introduction and as discussed in detail in the ISA, background ozone values

can vary significantly in space and time.  There can be atypical episodes of higher background ozone

concentrations amidst the routine days that drive seasonal average background.  The highest

background episodic concentrations are typically associated with stratospheric intrusions or wildfires.  

These background “events” can be difficult to model as they require event‐specific model inputs.  The

primary goal of the EPA modeling was to estimate the seasonal average background concentrations

between April and October 2007.  Previous analyses have shown that this is the period in which average

background levels are highest (Zhang et al., 2011).  This section of the appendix focuses on seasonal

mean levels of background.  (Section 4 will consider the upper range of possible background ozone.)   

The analysis focus on the maximum daily 8‐hour ozone average in ppb.  This metric is referred to

as MDA8.  This section will first present model estimates of seasonal mean ozone levels in the base

simulation.  This will be followed by estimates of NB, NAB, and USB from the CMAQ zero out modeling.  

After discussing the magnitudes of background levels, the section switches to a consideration of the

relative percentage of background to total ozone across the U.S.  This portion of the text will utilize both

the CMAQ zero out and CAMx source apportionment modeling.

Figure 3a displays the 2007 base case, CMAQ model‐predicted, seasonal mean (April‐October)

MDA8 ozone concentrations in grid cells with active monitoring locations over the U.S.  The model

results are shown at the monitoring site level as opposed to in the default gridded format to foster

subsequent site‐level estimates of background magnitudes.  Each grid cell containing an Air Quality

System (AQS) ozone monitor that was collecting valid data in 2007 was identified and the model
2A-15 

background estimates were extracted for those grid cells and displayed accordingly.  The base

predictions are provided for context to allow easier interpretation of the following plots which isolate

specific background levels.  As can be seen, most of the U.S. experiences seasonal mean MDA8 ozone

levels greater than 50 ppb in the base case simulation.  The median value over the 1,294 monitoring

locations is 52.5 ppb.

Figure 3b provides an estimate of what seasonal‐average MDA8 would be in a natural

background scenario, using the 2007 EPA zero out modeling.  Again, in this GEOS‐Chem/CMAQ

simulation, all anthropogenic ozone precursor emissions were removed from both the global and

regional simulations, and methane levels were adjusted to pre‐industrial levels in the global simulation.  

As shown, natural background ozone levels range from approximately 15‐35 ppb with the highest values

occurring over the higher‐elevation sites in the western U.S.  The median value over these locations is

24.2 ppb, and more than 50 percent of the sites have natural background levels of 20‐25 ppb.  The

highest modeled estimate of seasonal average, natural background, MDA8 ozone is 34.3 ppb at the

high‐elevation CASTNET site (Gothic) in Gunnison County, CO.

Figures 3c and 3d show the same information for the North American and U.S. background

scenarios.  In these model runs, all anthropogenic ozone precursor emissions were removed from the

U.S., Canada, and Mexico (NAB scenario) and then only the U.S. (USB scenario).  The figures show that

there is not a large difference between the NAB and USB scenarios.  Seasonal mean MDA8 NAB and USB

ozone levels range from 25‐50 ppb, with the most frequent values estimated in the 30‐35 ppb bin.  The

median seasonal mean background levels are 31.5 and 32.7 ppb (NAB and USB, respectively).  Again, the

highest levels of background are predicted over the intermountain western U.S.  Locations with NAB and

USB concentrations greater than 40 ppb are confined to Colorado, Nevada, Utah, Wyoming, northern

Arizona, eastern California, and parts of New Mexico.  Similar to NB, the highest NAB and USB levels

were modeled to occur at the Gothic CO site (46.7/47.7).  This remote rural site is located 2,926 meters

(9,600 feet) above mean sea level and should not be considered representative of background ozone at

lower‐altitude, more‐populated regions.  The high USB and NAB values along the Gulf Coast are most

likely due to model biases.

Absolute model estimates of various background definitions are useful, but they can be

influenced by any local biases and errors in the modeling.  A separate way to look at the role of

background in seasonal mean ozone levels is to consider the fractional contribution of NB, NAB, and USB

to total ozone at each location.  Considering the proportional role of background allows for an

informative comparison between the two modeling approaches without having to account for the

differences in base case biases and errors.  Figures 4a, 4b, and 4c show the estimated fractional

contribution of NB, NAB, and USB to total seasonal average MDA8 ozone levels at the monitoring

locations from the CMAQ zero out modeling.  The modeling estimates that approximately 35‐80 percent

of the seasonal average MDA8 ozone at monitoring locations is due to natural background sources.  A

majority sites have NB fractions between 40 and 60 percent.  The mean natural background proportion

over all sites is 47 percent.  That is, when all global anthropogenic emissions are removed and global

methane levels in GEOS‐Chem are restored to pre‐industrial levels, seasonal average MDA8 levels are

reduced by approximately half.   The fractional proportions of NAB and USB are very similar.  In both
2A-16 

cases, most sites have background fractions that range from 50 to 80 percent.  The mean NAB fraction

(to seasonal mean MDA8) is 63 percent.  The mean USB fraction is 66 percent.

As noted in the introduction, the advantage of the source apportionment modeling is that all of

the modeled ozone is attributed to various source terms and thus this approach is not affected by the

confounding occurrences of background ozone values exceeding the base ozone values as can happen in

the zero out modeling (i.e., background proportions > 100%).  Consequently, one would expect the

fractional background levels to be lower in the source apportionment methodology as a result of

removing this artifact.  It is also important to remember that the terms NB, NAB, and USB are explicitly

linked to the zero out modeling approach.  (USB is the ozone that would exist in the absence of U.S.

anthropogenic emissions.)  In contrast, the source apportionment modeling  performed here provides

estimates the amount of MDA8 ozone that is attributable to U.S. anthropogenic emissions  relative to

total base model ozone.  Figure 4d shows the relative contribution from sources other than U.S.

anthropogenic emissions to total seasonal mean MDA8 ozone based on the 2007 source apportionment

modeling.  The  fractional contribution fields between CMAQ zero out USB estimates and CAMx source

apportionment estimates of source other than U.S. anthropogenic emissions are quite similar.  The

spatial patterns in Figures 4c and 4d are consistent, with the highest fractional contributions from

sources other than U.S. anthropogenic emissions occurring along U.S. borders and over the

intermountain western States.  The source apportionment modeling estimates that approximately 40‐

80% of the seasonal average MDA8 ozone at monitoring locations is due to sources other than

manmade ozone precursor emissions from the U.S.  A majority of sites have non‐U.S. fractions between

40 and 70 percent.  The mean proportion attributable to international and natural sources over all sites

is 59 percent.  Despite the differences in the methodologies this is very similar to the mean USB

estimate of 66 percent from the zero out modeling.

   
2A-17 

Figure 3a.  April‐October average MDA8 ozone (ppb) at monitoring locations across the U.S. as

estimated by a 2007 CMAQ base simulation.

Figure 3b.  April‐October average natural background MDA8 ozone (ppb) at monitoring locations

across the U.S. as estimated by a 2007 CMAQ zero out simulation.
2A-18 

Figure 3c.  April‐October average North American background MDA8 ozone (ppb) at monitoring

locations across the U.S. as estimated by a 2007 CMAQ zero out simulation.

Figure 3d.  April‐October average United States background MDA8 ozone (ppb) at monitoring

locations across the U.S. as estimated by a 2007 CMAQ zero out simulation.   
2A-19 

Figure 4a.  Ratio of natural background to total April‐October average MDA8 ozone at monitoring

locations across the U.S. as estimated based on 2007 CMAQ simulations.

Figure 4b.  Ratio of N. American background to total April‐October average MDA8 ozone at

monitoring locations across the U.S. as estimated based on 2007 CMAQ simulations.
2A-20 

Figure 4c.  Ratio of U.S.  background to total April‐October average MDA8 ozone at monitoring

locations across the U.S. as estimated based on 2007 CMAQ simulations.

Figure 4d.  Ratio of sources other than U.S. anthropogenic emissions to total April‐October average

MDA8 ozone at monitoring locations across the U.S. as estimated by a 2007 CAMx source

apportionment simulation.
2A-21 

4. Distributions of background ozone levels

As a first‐order understanding, it is valuable to be able to characterize seasonal mean levels of

background ozone.  However, it is well established that background levels can vary substantially from

day‐to‐day.  From an implementation perspective, the values of background ozone on possible

exceedance days is a more meaningful distinction.  The first draft policy assessment (EPA, 2012)

considered this issue in detail, via summaries of the existing 2006 zero out modeling (Henderson et al.,

2012), and concluded that “results suggest that background concentrations on the days with the highest

total ozone concentrations are not dramatically higher than typical seasonal average background

concentrations.”  Based on this finding, the 1st draft policy assessment determined that “anthropogenic

sources within the U.S. are largely responsible for 4th highest 8‐hour daily maximum ozone

concentrations.”  This portion of the appendix will consider the entire spectrum of variable background

ozone levels with special emphasis on days in which base model ozone concentrations approach or

exceed the level of the NAAQS.   

The 2007 modeling agrees with the finding from the previous 2006‐based modeling analyses

that the highest modeled ozone site‐days tend to have background ozone levels similar to mid‐range

ozone days.  Figures 5a‐5c show the distribution of April‐October MDA8 background levels (NB, NAB,

and USB, respectively) from the CMAQ zero out runs.  As noted in section 2, zeroing out emissions can

remove the effects of local NOx titration and result in modeled background values that are higher than

the base model ozone. The “box and whisker” plots shown in these figures display four key features of

the distributions:  

a. the median concentration (black horizontal line) per bin,  

b. the inter‐quartile range (blue colored box) which represents the 25th‐75th percentile range in

values within the distribution,  

c. the “whiskers” (dark gray vertical lines with top and bottom whiskers) which represent the

range of values within 1.5 times the inter‐quartile range, and

d. the “outliers” (gray points) which are any values outside the whiskers.  

As can be seen in Figure 5a, natural background values do not vary greatly as a function of the

base modeled ozone.  Recall that the seasonal average natural background MDA8 ozone values were

modeled to range from 15‐35 ppb across the U.S. with a median value of 24 ppb.  The highest values

were at the high‐elevation sites in the western U.S.  Based on the distributional analysis, the 75th

percentile values are on the order of 30 ppb.  Natural background levels exceeding 40‐45 ppb are

considered to be statistical outliers, due to their infrequency.  Figure 5b shows the same type of

distributions but for NAB instead of NB.  NAB values are generally 6‐12 ppb higher than their NB

counterparts, due to the affect of higher global methane values and the influence of anthropogenic

emissions from Asia.  It was previously reported (in section 3) that the median seasonal average NAB

MDA8 values were 31.5 ppb.  Based on the distributions, it can be seen that 75th percentile values are
2A-22 

approximately 40 ppb; it is rare for NAB MDA8 values to exceed 50‐55 ppb.  NAB values are constant in

magnitude once the base ozone exceeds 50 ppb indicating that the higher base ozone values are driven

by non‐NAB sources (i.e., North American emissions).  Finally in Figure 5c, the USB MDA8 distributions

by base model MDA8 are shown.  The results are similar to NAB.

Figure 5d shows the results from the source apportionment modeling of non‐U.S. anthropogenic

source contributions to MDA8 ozone (i.e., the nine source apportionment categories other than U.S.

anthropogenic emissions and Gulf of Mexico point sources).  This non‐counterfactual approach is

expected to give a better indication of background levels at low concentrations.  At low levels, almost all

of the ozone is determined to be from background origins.  The CAMx modeling shows that

contributions from non‐U.S. anthropogenic emissions peak when base ozone ranges from 45‐55 ppb and

then drop off slightly at higher base MDA8 values.  The source apportionment modeling of non‐U.S.

impacts (similar to USB) indicates slightly lower background levels than the zero out modeling.  The 75th

percentile values are generally less than 35 ppb, compared to 40 ppb in the zero out modeling.  It is rare

to have background impacts greater than 55ppb.  Interestingly, when base model MDA8 ozone exceeds

70 ppb, it is rare to have background impacts greater than 45 ppb in the CAMx source apportionment

modeling.

Figures 6a‐6d show the equivalent plots as 5a‐5d, but use background fractions (background

MDA8 / base MDA8) as the dependent variable instead of the absolute background concentrations.  

These plots show the same effect; that is, the proportional relative contribution of background sources

and processes decreases as peak ozone increases.  For natural background (Figure 6a), the median

fractions drop from 50% background for values between 45‐50 ppb to only 35% background for base

MDA8 values between 70‐75 ppb.  For NAB and USB (Figures 6b and 6c), the median fractions drop from

70% background for values between 45‐50 ppb to only 45% background for base MDA8 values between

70‐75 ppb.  The source apportionment modeling (Figure 6d) estimates less of a proportional role of non‐

U.S. anthropogenic emissions.  In that modeling, the median fractions drop from 65% background for

values between 45‐50 ppb to only 35% background for base MDA8 values between 70‐75 ppb.  A key

observation, as noted in the first draft policy assessment document, is that the relative importance of

background decreases on days most likely to violate the NAAQS.  An additional policy‐relevant finding

from the distributional analyses is that the relative role of background sources would be increased if the

level of the NAAQS were lowered.  At 60 ppb, the modeling suggests that the median fractional

contribution from background is 45‐55 percent, but there can be cases where background comprises 80‐

90 percent of the total ozone.   

Many of the cases when background ozone is estimated to contribute in large proportions to

relatively high ozone days may be eligible for consideration as exceptional events, but again, this

modeling is not designed to resolve specific events that occurred in 2007.  While there is greater

confidence in the model’s ability to predict mean contributions from background sources than from

individual events, it is also useful to briefly consider the upper end of the background ozone
2A-23 

distributions.  Figure 7 shows the 95th percentile3 USB estimates from the zero out modeling.  The 95th

percentile MDA8 USB ozone levels range from 35‐60 ppb, with the most frequent values residing in the

35‐40 and 40‐45 ppb bins.  The median 95th percentile background USB ozone level is 42.0 ppb.  As with

the seasonal mean MDA8 USB, the highest levels of high background days (i.e., 95th percentile days) are

observed over the intermountain western U.S.  At these locations, 95th percentile USB levels can exceed

50 ppb.  Background values at the 95th percentile end of the distribution are 4‐12 ppb higher than the

mean background values at the same locations.

   

                                                            

3 During the April‐October period, there were 214 days of modeling results.  Thus, the 95th percentile values

represent approximately the 10th highest days from the distribution.
2A-24 

  

Figure 5a.  Distribution of natural background MDA8 ozone (ppb) at monitoring locations across the

U.S. (Apr‐Oct), binned by base modeled site‐day MDA8, as estimated by 2007 CMAQ simulations.

Figure 5b.  Distribution of N. American background MDA8 ozone (ppb) at monitoring locations across

the U.S. (Apr‐Oct), binned by base modeled site‐day MDA8, as estimated by 2007 CMAQ simulations.
2A-25 

Figure 5c.  Distribution of U.S. background MDA8 ozone (ppb) at monitoring locations across the U.S.

(Apr‐Oct), binned by base modeled site‐day MDA8, as estimated by 2007 CMAQ simulations.

Figure 5d.  Distribution of  MDA8 ozone contributions from non‐U.S. manmade sources (ppb) at

monitoring locations across the U.S. (Apr‐Oct), binned by base modeled site‐day MDA8, as estimated

by 2007 CAMx simulations.
2A-26 

Figure 6a.  Distribution of natural background MDA8 ozone fractions at monitoring locations across

the U.S. (Apr‐Oct), binned by base modeled site‐day MDA8, as estimated by 2007 CMAQ simulations.

Figure 6b.  Distribution of N. American background MDA8 ozone fractions at monitoring locations

across the U.S. (Apr‐Oct), binned by base modeled site‐day MDA8, as estimated by 2007 CMAQ

simulations.
2A-27 

Figure 6c.  Distribution of U.S. background MDA8 ozone fractions at monitoring locations across the

U.S. (Apr‐Oct), binned by base modeled site‐day MDA8, as estimated by 2007 CMAQ simulations.

Figure 6d.  Distribution of MDA8 ozone fractions from non‐U.S. anthropogenic sources at monitoring

locations across the U.S. (Apr‐Oct), binned by base modeled site‐day MDA8, as estimated by the 2007

CAMx simulation.
2A-28 

Figure 7.  April‐October 95th percentile United States background MDA8 ozone (ppb) at monitoring

locations across the U.S. as estimated by a 2007 CMAQ base simulation.

   
2A-29 

5. Contribution of various processes and sources to total background ozone  

This section will utilize the supplemental 2007 air quality modeling estimates to determine the

relative importance of specific elements of background ozone.  Comparing the differences between the

three zero out scenarios can provide some information about the role of certain sets of emissions.  

Figure 8a compares the NAB (zero out North American manmade emissions) and USB (zero out U.S.

manmade emissions) scenarios.  The difference between these two runs is the inclusion of

anthropogenic emissions within the Canada and Mexico portions of the modeling domain.  These

emissions contribute less than 2 ppb to the seasonal mean MDA8 ozone levels over most of the U.S.  

There are 70 sites, near an international border, where the modeling estimates Canadian/Mexican

seasonal average impacts of 2‐4 ppb.  While not shown, the modeled peak single day impacts from

these specific international emissions sources can approach 25 ppb (e.g., San Diego, Buffalo NY).  Figure

8b compares the NB (zero out all manmade emissions and reset GEOS‐Chem methane values to pre‐

industrial levels) to the NAB.  The difference between these two runs is the inclusion of global methane

emissions related to recent human activity as well as anthropogenic emissions outside of North America.  

These emissions are estimated to contribute 6‐15 ppb to seasonal mean ozone levels over the U.S.  The

most frequent bin is the 8‐10 ppb increase.  It is not possible via these runs to parse out what fraction of

this change is due to international emissions as opposed to methane emissions, but the ISA summarized

existing modeling (Zhang et al., 2011) that suggested that the rise in methane from pre‐industrial levels

to present‐day levels led to increases in seasonal average ozone levels of 4‐5 ppb.  The greatest impacts

from these sources occurs over the western U.S., where international emissions would be expected to

have the largest impacts.

Figures 9a‐9g show the fractional contribution to total seasonal mean MDA8 values of  

individual source sectors that were tracked in the CAMx source apportionment modeling.  Figure 9a

shows the impact from the regional model boundary conditions.  The ozone entering the model domain

via the boundary conditions could have a variety of origins including: a) natural sources of ozone and

ozone precursors (including methane) emanating from outside the domain, b) anthropogenic sources of

ozone precursors (including methane) from international emitters, and c) some fraction of U.S.

emissions (natural and anthropogenic) which are exported and then re‐imported into the domain via

synoptic‐scale recirculation.  Thus, one should not presume that the boundary condition contribution is

directly tied to any particular background definition.  At most locations, boundary conditions

contributed 40‐60 percent of the total MDA8 seasonal mean at sites across the U.S.  The highest

proportional impacts from the boundary conditions (the top boundary contributes negligibly) are along

the coastlines and the intermountain West.

Figure 9b shows the source apportionment contribution (to seasonal mean MDA8) from the

most significant sector that was tracked: U.S. anthropogenic ozone precursor emissions.  Again the most

common outcome at an individual site was that 40‐60% of the seasonal mean ozone values originated

from U.S. anthropogenic emissions.  The locations with smaller fractional contributions (e.g., 10‐20

percent) from U.S. sources are generally located in places where ozone values are typically low such as

the Pacific Northwest.  Figures 9c‐9g display the fractional contributions from the other five in‐domain

sectors listed in section 2.  The impacts from these sectors are briefly summarized below:
2A-30 

 Biogenic emissions:

o Most frequent bin: 3‐5 percent

o Highest site‐specific contribution: 10‐20 percent

o Region with greatest impacts: Great Plains states where soil NOx emissions are large  

 Climatologically‐average fire emissions:

o Most frequent bin: 0‐1 percent

o Highest site‐specific contribution: 3‐5 percent

o Region with greatest impacts: California, Kansas/Oklahoma region

 Within‐domain Canadian/Mexican manmade emissions:

o Most frequent bin: 0‐1 percent

o Highest site‐specific contribution: 10‐20 percent

o Region with greatest impacts: Sites along international borders (NY, VT, CA, AZ, TX)  

 Category 3 marine vessels outside U.S. territorial waters:

o Most frequent bin: 0‐1 percent

o Highest site‐specific contribution: 10‐20 percent

o Region with greatest impacts: Coastal sites (especially southern CA)

 Gulf of Mexico point sources4

:

o Most frequent bin: 0‐1 percent

o Highest site‐specific contribution: 1‐3 percent

o Region with greatest impacts: Sites in southeast TX and southern LA

   

                                                            

4

  This sector was also included as part of U.S. anthropogenic source impacts in Figure 9b, but is broken out

separately in Figure 9g.  
2A-31 

Figure 8a.  Difference in April‐October average MDA8 ozone (ppb) at monitoring locations across the

U.S. between the USB scenario and the NAB scenario.  The difference between these two runs isolates

the impact of within‐the‐domain anthropogenic emissions from Canada and Mexico.  

Figure 8b.  Difference in April‐October average MDA8 ozone (ppb) at monitoring locations across the

U.S. between the NAB scenario and the NB scenario.  The difference between these two runs isolates

the impact of the rise in global methane emissions from the pre‐industrial and anthropogenic

emissions from outside North America.  
2A-32 

Figure 9a.  Percentage of April‐October average MDA8 ozone that is apportioned to boundary

conditions  as estimated at monitoring locations by a 2007 CAMx simulation.

Figure 9b.  Percentage of April‐October average MDA8 ozone that is apportioned to U.S.

anthropogenic sources  as estimated at monitoring locations by a 2007 CAMx simulation.
2A-33 

Figure 9c.  Percentage of April‐October average MDA8 ozone that is apportioned to purely biogenic

emissions  as estimated at monitoring locations by a 2007 CAMx simulation.

Figure 9d.  Percentage of April‐October average MDA8 ozone that is apportioned to climatological fire

emissions  as estimated at monitoring locations by a 2007 CAMx simulation.
2A-34 

Figure 9e.  Percentage of April‐October average MDA8 ozone that is apportioned to anthropogenic

emissions from in‐domain Canadian and Mexican sources as estimated at monitoring locations by a

2007 CAMx simulation.

Figure 9f.  Percentage of April‐October average MDA8 ozone that is apportioned to Category 3 marine

vessel emissions beyond U.S. territorial waters as estimated at monitoring locations by a 2007 CAMx

simulation.

   
2A-35 

Figure 9g.  Percentage of April‐October average MDA8 ozone that is apportioned to Gulf of Mexico

point sources as estimated at monitoring locations by a 2007 CAMx simulation.

   
2A-36 

6. Estimates of the fractional background contribution to total ozone in 12 specific areas

This penultimate section of the appendix presents estimates of the overall fraction of ozone that

is estimated to result from background sources or processes based on the updated modeling in each of

the 12 urban case study areas in the epidemiology study based analyses in Chapter 7 of the Risk and

Exposure Assessment (REA).  Tables 1a‐1c summarize the CAMx‐estimated fractional contributions of

sources other than U.S. anthropogenic emissions to total ozone in each of the 12 areas.  Table 1a shows

that the fractional contributions from sources other than anthropogenic emissions within the U.S. to

seasonal mean MDA8 levels can range from 43 to 66 percent across these 12 urban areas.  These

fractions are consistent with the national ratios summarized in section 3, although the urban fractions of

background tend to be smaller than at rural sites.  As shown in section 4, the fractional contributions

from background are smaller on days with high modeled ozone (i.e., days that may exceed the level of

the NAAQS).  Table 1b provides the fractional contributions from these non‐U.S. sources, only

considering days in which base model MDA8 ozone was greater than 60 ppb.  As expected, the fractional

background contributions are less and range from 31 to 55 percent.  Rather than taking the fractions of

the seasonal means (as in Table 1a), Table 1c displays the mean and median daily MDA8 background

fractions.  These metrics may be more appropriate for application to health studies, but as can be seen

the fractional contribution to backgrounds calculated via this approach are very similar to the Table 1a

calculations.  For completeness sake, although EPA expects the source apportionment results to provide

a more realistic estimate of fractional background values, for completeness, we also provide USB

fractions based on zero out modeling for the 12 cities (see Table 1d).  The results are similar to the

source apportionment findings (compare against Table 1a), but the zero out technique provides slightly

higher background proportions.

Table 1a.  April‐October average MDA8 ozone, average MDA8 ozone from sources other than U.S.

manmade emissions, and the fractional contribution of these background sources in the 12 REA urban

study areas, as estimated by a 2007 CAMx simulation.

All days, CAMx ATL BAL BOS CLE DEN DET HOU LA NYC PHI SAC STL

Model MDA8 seasonal mean 59.3 54.4 43.0 48.9 47.3 39.1 48.5 51.1 45.4 48.7 46.4 49.8

Model MDA8 seasonal mean

from emissions other than

U.S. anthropogenic sources

25.3 25.9 26.2 25.7 31.3 23.3 27.0 29.1 24.5 24.2 29.7 24.3

Fractional contribution from

background 0.43 0.48 0.61 0.52 0.66 0.60 0.56 0.57 0.54 0.50 0.64 0.49
2A-37 

Table 1b.  Average MDA8 ozone, average MDA8 ozone from sources other than U.S. manmade

emissions, and the fractional contribution of these background sources in the 12 REA areas, as

estimated by a 2007 CAMx simulation using site‐days in which base MDA8 ozone exceeded 60 ppb.

Table 1c.  Fractional contribution of non‐U.S. manmade emissions sources in the 12 REA urban study

areas, as estimated by a 2007 CAMx simulation using means and medians of daily MDA8 fractions.  

Table 1d.  April‐October average MDA8 ozone, average MDA8 ozone from USB, and the fractional

contribution of these background sources in the 12 REA urban study areas, as estimated by two

separate 2007 CMAQ simulations.

Only days w/ base

MDA8 > 60 ppb ATL BAL BOS CLE DEN DET HOU LA NYC PHI SAC STL

Model MDA8 seasonal mean 74.0 75.3 70.7 72.0 67.5 68.9 70.3 74.4 74.1 74.0 68.3 70.0

Model MDA8 seasonal mean

from emissions other than

U.S. anthropogenic sources

25.4 23.7 24.4 25.4 37.3 24.4 28.0 31.9 23.5 22.9 32.1 25.4

Fractional contribution from

background 0.34 0.31 0.35 0.35 0.55 0.35 0.40 0.43 0.32 0.31 0.47 0.36

ATL BAL BOS CLE DEN DET HOU LA NYC PHI SAC STL

Mean of daily MDA8

background fractions 0.46 0.53 0.68 0.58 0.69 0.64 0.59 0.61 0.61 0.56 0.67 0.52

Median of daily MDA8

background fractions 0.43 0.51 0.73 0.54 0.69 0.66 0.59 0.60 0.63 0.54 0.66 0.49

All days, CMAQ ATL BAL BOS CLE DEN DET HOU LA NYC PHI SAC STL

Model MDA8 seasonal mean 58.6 55.6 45.2 51.8 57.1 43.5 49.4 54.8 47.7 50.5 51.9 52.6

Model MDA8 seasonal mean

from USB emissions

30.0 29.9 28.5 31.6 42.2 31.7 33.0 33.3 29.1 29.4 34.4 32.0

Fractional contribution from

background 0.51 0.54 0.63 0.61 0.74 0.73 0.67 0.61 0.61 0.58 0.66 0.61
2A-38 

7. Background ozone and W126

As discussed in section 5 of the second draft policy assessment, EPA is considering the adequacy

of the current secondary standard to protect against welfare effects.  One metric that has been

considered previously as a potential cumulative seasonal index is the W126 metric.  The W126 index is a

sigmoidally weighted sum of all hourly O3 concentrations observed during a specified daily and seasonal

time window, where each hourly O3 concentration is given a weight that increases from 0 to 1 with

increasing concentration (Lefohn et al, 1988).  The weights are defined such that values of 0.060 ppm

get a weight of ~0.3; 0.070 ppm values get a weight of ~0.6; and 0.085 ppm values get a weight of ~0.9.  

The remainder of this section uses the 2007 zero out modeling to conduct a limited assessment of the

role of background ozone on W126 levels over the U.S.

The analysis of background influence on W126 is not as detailed as the analyses related to

seasonal mean MDA8 ozone.  Instead of considering impacts at every monitoring location, EPA assessed

NB, NAB, and USB influences at four sample locations: Atlanta GA, Denver CO, Farmington NM, and

Riverside CA.  Each of these four locations had relatively high observed values of W126 in 2010‐2012.  

Atlanta is an urban area in the Eastern U.S. with high primary ozone design values but relatively low

levels of seasonal background ozone.  Riverside and Denver also have high primary ozone design values

but are in the Western U.S. where background ozone levels are generally higher.  Farmington NM was

chosen as a site that has relatively lower primary ozone design values along with its relatively high W126

levels.  The varying characteristics of each of these locations perhaps allows broader national

extrapolation of the 4‐site results.

In previous EPA reviews of the O3 NAAQS, the influence of background ozone was estimated

according to a counterfactual (i.e., how much ozone would exist in the absence of certain sets of

emissions).  In the current review, EPA is supplementing the counterfactual assessment with analyses

that estimate the fraction of the existing ozone that is due to background sources.  This has important

ramifications for assessing the influence of background on W126 concentrations, because of the non‐

linear weighting function used in the metric which emphasizes high ozone hours (e.g., periods in which

ozone is greater than ~60 ppb).  As an example, consider a sample site in the intermountain western

U.S. region with very high modeled estimates of U.S background (e.g., seasonal mean USB of 45 ppb

with some days as high as 65 ppb).  Even at this high background location, the calculated annual W126

values in the USB scenario are quite low, on the order of 3 ppm‐hrs.  Most sites in the domain where

background levels are lower than the location cited above will have even smaller background W126

estimates, on the order of 1 ppm‐hrs, which is consistent with values mentioned in past reviews (USEPA,

2007).  Using the counterfactual scenarios, background ozone has a relatively small impact on W126

levels across the U.S.

However, because of the non‐linear weighting function used in the W126 calculation, the sum of

the W126 from the USB scenario and the W126 resulting from US anthropogenic sources will not equal

the total W126.  In most cases, the sum of those two components will be substantially less than total

W126.  As a result, EPA believes it is more informative to estimate the fractional contribution of

background ozone to W126 levels.  The 5‐step methodology for assessing the fractional influence of
2A-39 

background ozone to annual W126 levels in the four locations is described below.  The fractional

influence methodology essentially places higher weights on background fractions on days that are going

to contribute most substantially to the yearly W126 value.

 Step 1a: Calculate the MDA8 ozone values from the base and the three zero out

modeling scenarios at each grid cell containing a site in an area.

 Step 1b: Calculate the W126 daily index for the base model scenario.

 Step 2: For each site, find the three months with highest summed W126 daily indices.

 Step 3: Normalize the daily MDA8 values in the base, NB, NAB, and USB scenarios by the

corresponding W126 daily index from the base scenario.

 Step 4: Calculate the average W126‐weighted MDA8 values over the three month

period for each of the four scenarios (base, NB, NAB, USB).

 Step 5: Calculate the NB/Base, NAB/Base, and USB/Base ratios based on step 4 outputs.  

These values represent an estimate of the fractional influence of background ozone on

modeled W126 levels.  

Figure 7a shows the estimated fractional influence of the three background definitions on W126

levels in Atlanta, Denver, Farmington, and Riverside.  Based on this limited assessment, natural

background sources are estimated to contribute 29‐50% of the total modeled W126 with the highest

relative influence in the intermountain western U.S. (e.g., Farmington NM) and the lowest relative

influence in the eastern U.S. (e.g., Atlanta).  U.S. background is estimated to contribute 37‐65% of the

total modeled W126.  Figure 7b compares the relative influence of background on W126 versus seasonal

mean MDA8 ozone.  The proportional impacts of background are slightly less for the W126 metric than

for seasonal mean MDA8 (discussed in section 2.4.2), because of the weighting function that places

more emphasis on higher ozone days when background fractions are generally lower.

There are several caveats associated with this analysis.  First, only the zero out modeling was

used to assess the fractional influence of background sources on W126.  The source apportionment

approach estimated slightly smaller relative contributions for seasonal mean MDA8 levels, so from that

perspective the zero out estimates could represent the high end of background influence on W126.  

Additionally, the methodology used for this analysis relies on daily MDA8 values as a surrogate (the data

were readily available) for the 8a‐8p time period relevant to the W126 metric.  The key conclusion from

this cursory analysis is that background ozone may comprise a non‐negligible portion of current W126

levels across the U.S.  This fractional influence is greatest in the intermountain western U.S. and are

slightly smaller than the seasonal mean MDA8 metric.  In the counterfactual cases, when non

background sources are completely removed, the remaining W126 levels are low (< 3 ppm‐hrs).
2A-40 

Figure 7a.  Fractional contribution of background sources to W126 levels in four sample locations.  

Model estimates based on 2007 CMAQ zero out modeling.

Figure 7b.  Fractional contribution of U.S. background to seasonal mean MDA8 ozone and W126 levels

in four sample locations.  Model estimates based on 2007 CMAQ zero out modeling.    

0%

10%

20%

30%

40%

50%

60%

70%

80%

90%

100%

Farmington Denver Riverside Atlanta

Percent contribution from background to W126

NB

NAB

USB

0%

10%

20%

30%

40%

50%

60%

70%

80%

90%

100%

Farmington Denver Riverside Atlanta

Percent contribution from backgrond

MDA8

W126
2A-41 

8. Summary  

The precise definition of background ozone can vary depending upon context, but it generally

refers to ozone that is formed by sources or processes that cannot be influenced by local control

measures.  Background ozone can originate from natural sources of ozone and ozone precursors, as well

as from upwind manmade emissions of ozone precursors.  In order to help further characterize

background ozone levels over the U.S., EPA has completed additional air quality modeling analyses

subsequent to the 1st‐draft policy assessment.  As shown above, the results are largely consistent with

previous determinations about the magnitude of background ozone contributions across the U.S.

For a variety of reasons, it is challenging to present a comprehensive summary of all the

components and implications of background ozone.  In many forums the term “background” is used

generically and the lack of specificity can lead to confusion as to what sources are being considered.  

Additionally, it is well established that the impacts of background sources can vary greatly over space

and time which makes it difficult to present a simple summary of background ozone levels.  Further,

background ozone can be generated by a variety of processes, each of which can lead to differential

patterns in space and time, and which often have different regulatory ramifications.  Finally, background

ozone is difficult to measure and thus, typically requires air quality modeling which has inherent

uncertainties and potential errors and biases.  Even with all of these complexities in mind, EPA believes

the following concise and step‐wise summary of background ozone is appropriate as based on previous

modeling exercises and the more recent EPA analyses summarized herein.

 The most fundamental definition of background is “natural background” (NB).  NB ozone is that

which is produced by processes other than manmade emissions.  Examples of sources of natural

background include: stratospheric ozone intrusions, wildfire emissions, and biogenic emissions

from vegetation and soils.  To date, NB ozone has been estimated to be that ozone that would

exist in the absence of anthropogenic ozone precursor emissions worldwide.  Modeling analyses

have shown that NB levels can vary in time and space.  As shown in Section 3, April‐October

average NB levels range from approximately 15‐35 ppb with the highest values in the spring and

at higher‐elevation sites.

 More expansive definitions of background include North American background (NAB) and U.S.

background (USB).  These definitions represent the ozone that originates from sources and

processes other than North American or U.S. anthropogenic sources.  Sources of NAB and USB

include all the same sources of natural background, plus manmade ozone precursors emitted

outside the North America or the U.S.  Modeling analyses have shown that NAB and USB

background levels can vary in time and space.  As discussed in Section 3, seasonal mean NAB

and USB background levels range from approximately 25‐45 ppb with the highest values in the

spring and at higher‐elevation sites.  USB levels are slightly higher than NAB, usually by less than

2 ppb.
2A-42 

 Estimates of seasonal mean background ozone levels are valuable in terms of a first‐order

characterization, however because levels can vary significantly from day‐to‐day, it is also

instructive to consider the distribution of daily model estimates of background ozone over a

season.  Typically, model background is slightly higher in the April‐June period than in the later

portion of the ozone season (July‐October) (EPA, 2012).  More importantly, the modeling shows

that the days with highest ozone levels, on average, have similar background levels to days with

lower values.  As a result, the proportion of total ozone that has background origins is smaller

on high ozone days (e.g., days > 70 ppb) than the more common lower ozone days that drive

seasonal means.  Section 4 provides information about the distribution of background ozone

fractions.  Based on the source apportionment modeling, it is shown that U.S. anthropogenic

emissions typically comprise the majority of the total ozone on site‐days with base modeled

ozone MDA8 values greater than 60 ppb.

 While it is important to recognize that most high ozone days (i.e., potential exceedance days)

are estimated to be driven predominantly by non‐background emissions, the recent EPA

modeling also shows times and locations in which background contributions are estimated to

approach 60‐80 ppb.  As described in Sections 4 and 6 of this document, these occurrences are

relatively infrequent.  While the modeling was not expressly developed to capture these types

of events, ambient observations have also shown relatively rare events where background

ozone sources (wildfires, stratospheric intrusions) have overwhelmingly contributed to an

ozone exceedance.  From a policy perspective, these background events must be viewed in the

context of their relative infrequency and the existing mechanisms within the Clean Air Act (e.g.,

exceptional event policy, 179B international determinations) that help ensure States are not

required to control for events that are inherently outside their ability to influence.   While

background ozone levels can approach and periodically exceed the NAAQS at some locations,

these conditions are not a constraining factor in the selection of a NAAQS.  The Clean Air Act

requires the NAAQS to be set at a level requisite to protect public health and welfare.  Case law

makes it clear that attainability and technical feasibility are not relevant considerations.  In

previous reviews, EPA assessed the proximity of potential levels to peak background levels as a

secondary consideration between levels where health and welfare was protected.

 Section 5 shows that the contributions to background are multi‐dimensional.  Daily peak 8‐hour

ozone values over the U.S. are a function of local and regional anthropogenic emissions,

anthropogenic emissions from outside the U.S. (including shipping emissions), natural and

anthropogenic methane emissions,  wildfire emissions, and purely natural sources.   While local

and regional controls are still considered to be the most effective at reducing local ozone levels,

any measures to reduce the international contributions or methane‐induced background will

also be valuable.

 In previous ozone NAAQS reviews, EPA estimated risk from exposure only to ozone

concentrations above background.  In the first drafts of the REA and PA for the current ozone
2A-43 

review, EPA estimated risk from exposure to total measured ozone concentrations, which

include those concentrations from background sources.  EPA will continue to provide estimates

of risk from exposure to total ozone, consistent with CASAC advice, in the second draft policy

assessment.  The recent EPA modeling was completed to assist in determining, in a limited

sense, the risk attributable to background ozone.  The fractional values of background

contributions in the 12 REA study areas (43‐66 percent) could be used as first order

approximations of the risk due to ozone background.   

9. References  

Arunachalam, S.  (2009). Peer Review of Source Apportionment Tools in CAMx and CMAQ.  Prepared

under EPA Contract #EP‐D‐07‐102, Institute of the Environment, University of North Carolina,

Chapel Hill, NC, 42pp,

http://www.epa.gov/scram001/reports/SourceApportionmentPeerReview.pdf.

Emery, C; Jung, J; Downey, N; Johnson, J; Jimenez, M; Yarwood, G; Morris, R. (2012). Regional and global

modeling estimates of policy relevant background ozone over the United States. Atmospheric  

Environment, 47: 206‐217. http://dx.doi.org/10.1016/j.atmosenv.2011.11.012.

Environ (2011).  User’s Guide: Comprehensive Air Quality Model with Extensions, Version 5.40; Novato

CA, 306pp.  http://www.camx.com/files/camxusersguide_v5‐40.aspx.

Fiore, A; Jacob, DJ; Liu, H; Yantosca, RM; Fairlie, TD; Li, Q. (2003). Variability in surface ozone background

over the United States: Implications for air quality policy. J Geophys Res, 108: 4787.

http://dx.doi.org/10.1029/2003JD003855.

Henderson, BH; Possiel N; Akhtar F; Simon H. (2012). Memo to the Ozone NAAQS Review Docket EPA‐

HQ‐OAR‐2012‐0699: Regional and Seasonal Analysis of North American Background Ozone

Estimates from Two Studies.  

http://www.epa.gov/ttn/naaqs/standards/ozone/s_o3_2008_td.html.

Lefohn, A. S.; Laurence, J. A.; Kohut, R. J. (1988). A comparison of indices that describe the relationship

between exposure to ozone and reduction in the yield of agricultural crops. Atmos. Environ. 22:

1229‐1240.  

Simon, H., Baker, K.P., Phillips, S. (2012) Compilation and interpretation of photochemical model

performance statistics published between 2006 and 2012.  Atmospheric Environment, 61, 124‐

139.

U.S. Environmental Protection Agency (2005).  Technical Support Document for the Final Clean Air

Interstate Rule Air Quality Modeling.  Office of Air Quality Planning and Standards, Research

Triangle Park, NC, 285pp.  http://www.epa.gov/cair/technical.html.

U.S. Environmental Protection Agency. (2007). Review of the National Ambient Air Quality Standards for

Ozone: Policy Assessment of Scientific and Technical Information ‐  OAQPS Staff Paper, U.S.

Environmental Protection Agency, Research Triangle Park, NC. EPA 452/R‐07‐007.
2A-44 

U.S. Environmental Protection Agency (2011).  Air Quality Modeling Final Rule Technical Support

Document.  Office of Air Quality Planning and Standards, Research Triangle Park, NC, 363pp.  

http://www.epa.gov/airtransport/CSAPR/techinfo.html.

U.S. Environmental Protection Agency (2012).  Policy Assessment for the Review of the Ozone National

Ambient Air Quality Standards: First External Review Draft.  EPA‐452/P‐12‐002, 297pp.

U.S. Environmental Protection Agency. (2013). Integrated Science Assessment for Ozone and Related

Photochemical Oxidants, U.S. Environmental Protection Agency, Research Triangle Park, NC.

EPA/600/R‐10/076.

U.S. Environmental Protection Agency. (2014). Health Risk and Exposure Assessment for Ozone, Second

External Review Draft, U.S. Environmental Protection Agency, Research Triangle Park, NC. EPA

xxx/P‐xx‐xxx.

Zhang, L; Jacob, DJ; Downey, NV; Wood, DA; Blewitt, D; Carouge, CC; Van donkelaar, A; Jones, DBA;

Murray, LT; Wang, Y. (2011). Improved estimate of the policy‐relevant background ozone in the

United States using the GEOS‐Chem global model with 1/2 2/3 horizontal resolution over North

America. Atmospheric Environment 45: 6769‐6776.

http://dx.doi.org/10.1016/j.atmosenv.2011.07.054.  
2B-1   

APPENDIX 2B 

MONITORING DATA ANALYSIS OF RELATIONSHIPS 

BETWEEN CURRENT STANDARD AND W126 METRIC 

Presented here are monitoring data analyses evaluating relationship between ozone (O3) 

concentrations in the averaging time and form of the current secondary standard (3-year average 

of the annual 4th highest daily maximum 8-hour concentrations, in parts per billion), and a threeyear W126 metric (3-year average of the annual maximum 3-month sum of weighted daytime 

concentrations, in parts per million-hours). We also consider the responsiveness of these two 

metrics to historical changes in air quality related to ozone precursor emissions. 

For this analysis, we chose to examine monitoring data from a base period (2001-2003) 

as well as a recent period (2009-2011). The base period was chosen to represent air quality 

conditions before the implementation of the 1997 national ambient air quality standard 

(NAAQS) for O3 (0.08 ppm). In 2004, EPA designated 113 areas as nonattainment for the 1997 

standard, which required many areas to begin precursor emissions control programs for the first 

time. At about the same time, EPA began implementation of the NOx Budget Trading Program 

under the NOx State Implementation Plan, also known as the “NOx SIP Call1

,” which required 

summertime reductions in NOx emissions from power plants and other large sources throughout 

the Eastern U.S. These programs were successful in reducing peak O3 concentrations, especially 

in the Eastern U.S., and as a result only 8 of the original 113 nonattainment areas were still 

violating the 1997 O3 NAAQS during the 2009-2011 period. 

Hourly O3 concentration data were retrieved from EPA’s Air Quality System (AQS) 

database2 for both periods, and used to calculate design values for the current standard as well as 

3-year average W126 values for both periods. The procedures for calculating design values for 

the current standard from hourly O3 concentration data are described in 40 CFR Part 50, 

Appendix P, and the procedures for calculating the 3-year average W126 values are described in 

section 4.3.1. of the 2nd draft Welfare Risk and Exposure Assessment (WREA). There were 838 

monitoring sites with sufficient data to calculate these values for both periods. In order to 

identify regional patterns in the relationships, these sites were grouped into the nine NOAA 

                                                            

1

 http://www.epa.gov/airmarkets/progsregs/nox/sip.html 

2

 EPA’s Air Quality System (AQS) database is a national repository for many types of air quality and 

related monitoring data. AQS contains monitoring data for the six criteria pollutants dating back to the 1970’s, as 

well as more recent additions such as PM2.5 speciation, air toxics, and meteorology data. At present, AQS receives 

hourly O3 monitoring data collected from nearly 1,400 monitors operated by over 100 state, local, and tribal air 

quality monitoring agencies. 
2B-2   

climate regions (Karl and Koss, 1984) used in the WREA. Figure 2B-1 presents a map of these 

regions, which are color-coded to match the scatter plots in the subsequent figures. 

Figures 2B-2a, 2B-2b, 2B-3a and 2B-3b show scatter plots of the design values for the 

current standard (x-axis) versus 3-year average W126 values (y-axis) for the base period and 

recent period, respectively. Most monitors in the U.S. both exceeded the current standard of 75 

ppb and a three-year average W126 value of 15 ppm-hrs during the base period. During the 

recent period, both the design values and 3-year average W126 values were much lower, and 

there also appears to be less scatter between the two metrics. In both periods, the highest design 

values and W126 values occurred in the West region which includes California. Finally, it is 

worth noting that monitors in the Southwest and West regions tend to have higher W126 values 

relative to their design values than in other regions. 

Figure 2B-4 shows a scatter plot of the design values for the current standard for the base 

period (x-axis) versus for the recent period (y-axis), while Figure 2B-5 shows this same 

relationship based on the 3-year average W126 values. The relationship between the two periods 

appears to be fairly linear for both metrics, indicating that larger decreases in these metrics 

tended to occur at monitors with higher base values. Figures 2B-6 and 2B-7 show design values 

for the current standard and 3-year average W126 values, respectively, compared to the unit 

changes in those values between the base period and recent period. Figures 2B-6 and 2B-7 show 

the difference between each point and the one-to-one lines in Figures 2B-4 and 2B-5, 

respectively. In particular, these figures highlight that there were some monitors where design 

values for the current standard and/or W126 values increased. However, those monitors also 

tended to have lower base values, and were mostly located outside of areas subject to emissions 

controls under the 1997 standard. 

Finally, Figure 2B-8 compares the unit change in design values (in ppb; x-axis) to the 

unit change in 3-year average W126 values (in ppm-hrs; y-axis). This figure shows that in most 

locations, the current standard metric and the W126 metric exhibit similar responses to changes 

in precursor emissions. In particular, the NOx SIP Call, which was implemented in the states 

east of the Mississippi River, was effective at reducing both design values and W126 values at 

nearly all monitors in the Eastern U.S. The relationship was much more variable in the 

remaining regions, where emissions control programs were mostly local and limited to areas 

which were violating the NAAQS. 

Based on this analysis of ambient monitoring data, we can make the following general 

conclusions about the relationship between the design value metric for the current O3 standard 

and the 3-year average W126 metric: 

1. There is a fairly strong, positive degree of correlation between the two metrics. 
2B-3   

2. Monitors in the West and Southwest regions tend to have higher W126 values relative to 

their design values than in other regions. 

3. Reducing precursor emissions, especially NOx, is an effective strategy for lowering both 

design values and W126 values. In particular, regional control programs such as the NOx 

SIP call are effective at reducing both metrics over a broad area. 

In addition, Figure 2B-9 examines the number of counties with 8-hour design values 

meeting the current standard and 3-year average W126 index values greater than 15 ppm-hrs. 

Most of these counties were located in the Southwest region of the country. There were no 

counties in any of the studied 3-year periods that had design values less than or equal to 65 ppb 

and 3-year average W126 index values greater than 15 ppm-hrs. 

Figure 2B-1. Map of the 9 NOAA climate regions (Karl and Koss, 1984), color coded to 

match the subsequent scatter plots. 
2B-4   

Figure 2B-2a. Design values for the current O3 standard in ppb (x-axis) versus 3-year 

average W126 values in ppm-hrs (y-axis) based on ambient monitoring data 

for 2001-2003. 
2B-5   

Figure 2B-2b. Design values for the current O3 standard in ppb (x-axis) versus 3-year 

average W126 values in ppm-hrs (y-axis) based on ambient monitoring data 

for 2001-2003 with a focus on monitors with 2001-2003 design values below 75 

ppb. 
2B-6   

Figure 2B-3a. Design values for the current O3 standard in ppb (x-axis) versus 3-year 

average W126 values in ppm-hrs (y-axis) based on ambient monitoring data 

for 2009-2011. 
2B-7   

Figure 2B-3b. Design values for the current O3 standard in ppb (x-axis) versus 3-year 

average W126 values in ppm-hrs (y-axis) based on ambient monitoring data 

for 2009-2011 with a focus on monitors with 2009-2011 design values below 75 

ppb. 
2B-8   

Figure 2B-4. Design values for the current O3 standard in ppb based on ambient 

monitoring data for 2001-2003 (x-axis) versus 2009-2011 (y-axis). 
2B-9   

Figure 2B-5. Three-year average W126 values in ppm-hrs based on ambient monitoring 

data for 2001-2003 (x-axis) versus 2009-2011 (y-axis). 
2B-10   

Figure 2B-6. Design values for the current O3 standard in ppb based on ambient 

monitoring data for 2001-2003 (x-axis) versus unit (ppb) change in design 

values from 2001-2003 to 2009-2011 (y-axis). 
2B-11   

Figure 2B-7. Three-year average W126 values in ppm-hrs based on ambient monitoring 

data for 2001-2003 (x-axis) versus unit (ppm-hr) change in 3-year average 

W126 values from 2001-2003 to 2009-2011 (y-axis). 
2B-12   

Figure 2B-8. Unit (ppb) change in design values for the current O3 standard from 2001-

2003 to 2009-2011 (x-axis) versus unit (ppm-hr) change in 3-year average 

W126 values from 2001-2003 to 2009-2011 (y-axis). 
January 2014 2B-13 Draft - Do Not Quote or C

1

2

3 Figure 2B-9. Number of counties where the 8-hour design value is meeting the current standard and 3-yea4 index value is greater than 15 ppm-hrs (left), and number of counties where the 8-hour desig

or equal to 70 ppb and 3-year average W126 index value is greater than15 ppm-hrs (right)3 5 . 

67 REFERENCES 8 Karl, T.R. and Koss, W.J., 1984: "Regional and National Monthly, Seasonal, and Annual Temperature Weighted by Area, 1895-1983." 9 Series 4-3, National Climatic Data Center, Asheville, NC, 38 pp. 

3 No counties in any of the studied 3-year periods were at or below a 3-year average of 4th highest daily maximum 8-hour av

above a 3-year W126 index value of 15 ppm-hrs. 
2C-1 

APPENDIX 2C 

INTER-ANNUAL VARIABILITY IN W126 INDEX VALUES: 

COMPARING ANNUAL AND 3-YEAR AVERAGE METRICS (2008-2010) 

2C.1 OVERVIEW 

This appendix describes an analysis comparing values for a single-year or annual W126 

metric to a W126 metric averaged over three consecutive years. The purpose of this analysis is 

to compare values based on a 3-year average of annual W126 indices to values based on a single 

annual W126 index. The deviations of the annual W126 index values in 2008, 2009, and 2010 

from the 2008-2010 average W126 index values are presented. 

2C.2 GENERAL DATA PROCESSING 

The air quality data for this analysis originated from EPA’s Air Quality System (AQS) 

data base, the official repository of ambient air measurements. The data used in this analysis 

consisted of W126 index values calculated from hourly ozone concentrations measured at 1082 

ozone monitors nationwide. Ozone monitors must have submitted data to AQS for at least 75% 

days in their required ozone monitoring season in 2008, 2009, and 2010 to be included in the 

analysis. 

2C.3 RESULTS & CONCLUSION 

The figure below shows a scatter plot of the deviations in the annual W126 index from 

the 3-year average by monitor. The solid curves represent the average deviation in a moving 

window along the x-axis for each year. From this figure, it is apparent that the highest annual 

W126 index value occurred in 2008 for most monitoring locations, the lowest annual W126 

index value occurred in 2009 for most monitoring locations, and the 2010 W126 index value was 

generally somewhere in between. It is also apparent that the inter-annual variability in the W126 

index increases along with the 3-year average. For monitors with 3-year average W126 values 

near 15 ppm-hrs, the average deviation was +3.5 ppm-hrs in 2008 and -3.8 ppm-hrs in 2009. 

This represents a 1-year swing of -7.3 ppm-hrs. 

The model-based air quality adjustments in the 2nd draft of the O3 Welfare REA show 

that reducing NOx emissions is effective for reducing 3-year average W126 levels. In Appendix 

2B, the analyses based on ambient monitoring data also show that large-scale reductions in NOx 

emissions are associated with lower W126 levels. Finally, the data analysis presented in this 

appendix shows that the inter-annual variability in the annual W126 index tends to decrease with 

decreasing W126 levels. Thus, it is expected that reductions in NOx emissions will not only 
2C-2 

result in lower 3-year average W126 levels, but also result in less inter-annual variability 

associated with annual W126 levels. 

The W126 index is based on a logistic weighting function that increases the weights 

assigned to hourly ozone concentrations very rapidly. Hourly ozone concentrations of 50 parts 

per billion are given a weight of about 10% while concentrations of 80 parts per billion are given 

a weight of nearly 90%. The annual W126 index is calculated as a 3-month sum of weighted 

ozone concentrations during daylight hours, which amounts to a sum of roughly 1100 weighted 

hourly concentrations. Thus, even a modest change in the average daily ozone level may have a 

significant impact upon the annual W126 index. Since ozone formation is heavily influenced by 

meteorology, the inter-annual variability in meteorological conditions tends to cause a large 

inter-annual variability in the W126 index. 

In conclusion, this evaluation indicates the extent to which a form for the secondary 

ozone standard that averages the annual W126 index values over three consecutive years might 

be expected to account for the annual variability in this index since the 3-year period would be 

expected to include year(s) below as well as above the 3-year average. 

 
2C-3 

Figure 2C-1. Deviation of the annual W126 index values in 2008, 2009, and 2010 (y-axis) 

from the 3-year average W126 index value (x-axis). 
3A-1 

APPENDIX 3A 

RECENT STUDIES OF RESPIRATORY-RELATED 

EMERGENCY DEPARTMENT VISITS AND HOSPITAL 

ADMISSIONS 

Hospital Admissions for All Respiratory Causes 

The APHENA study (APHENA is for Air Pollution and Health: A European and North 

American Approach) analyzed air pollution and health outcome data from existing Canadian, 

European, and U.S. multi-city studies and examined the influence of varying model specification 

to control for season and weather (Katsouyanni et al., 2009). The U.S.-based portion of the 

APHENA study utilized the National Morbidity, Mortality, and Air Pollution Study (NMMAPS) 

cohort which, for the Katsouyanni et al. (2009) analysis, comprised respiratory hospital 

admissions among individuals 65 years of age and older from 14 US cities with O3 data from 

1985-1994 (7 cities had summer only O3 data). For the year round analysis, Katsouyanni et al. 

(2009) reported consistently positive, and statistically significant in models with 8 degrees of 

freedom per year (U.S. EPA, 2013, section 6.2.7.2), associations between 1-hour O3

concentrations and respiratory hospital admissions across the datasets from the U.S., Canada, and 

Europe (U.S. EPA 2013, Figure 6-15).1

 In co-pollutant models adjusting for PM10, O3 effect 

estimates remained positive, though effect estimates were somewhat attenuated in the U.S. and 

European datasets, possibly due to the PM sampling schedule (U.S. EPA 2013, Figure 6-15). 

Effect estimates for the warm season were larger than for the year-round analysis in the 

Canadian dataset, but generally similar in magnitude to the year-round analysis in the U.S. and 

European datasets. 

Several additional multicity studies examined respiratory disease hospital admissions in 

Canada and Europe. Cakmak et al. (2006) reported a statistically significant increase in 

respiratory hospital admissions in 10 Canadian cities (4.4% increase per 20 ppb increase in 24-

hour average O3, 95% CI: 2.2, 6.5%). In analyses of potential effect modifiers of the O3-

respiratory hospital admission relationship, individuals with an education level less than the 9th

grade were found to be at greater risk. Dales et al. (2006) reported a 5.4% (95% CI: 2.9, 8.0%) 

increase in neonatal respiratory hospital admissions for a 20 ppb increase in 24-hour average O3

 1

The study by Katsouyanni et al. (2009) evaluated different statistical models. Although the investigators 

did not identify the model they deemed to be the most appropriate for comparing the results across study locations, 

they did specify that “overall effect estimates (i.e., estimates pooled over several cities) tended to stabilize at high 

degrees of freedom” (Katsouyanni et al., 2009). In discussing of the results of this study, the ISA focused on models 

with 8 degrees of freedom per year (US EPA, 2012a, section 6.2.7.2). 
3A-2 

concentrations in 11 Canadian cities from 1986 to 2000. In contrast, Biggeri et al. (2005) did not 

detect an association between short-term O3 exposure and respiratory hospital admissions in four 

Italian cities from 1990 to 1999. 

In addition to the large multi-city studies discussed above, several smaller-scale studies 

have also reported associations with total respiratory hospital admissions. Specifically, Lin et al. 

(2008) reported a positive association between O3 and pediatric (i.e., <18 years) respiratory 

admissions in an analysis of 11 geographic regions in New York state from 1991 to 2001, though 

results were not presented quantitatively. In co-pollutant models with PM10, the authors reported 

that region-specific O3 associations with respiratory hospital admissions remained relatively 

robust. 

Cause-Specific Hospital Admissions 

With regard to cause-specific respiratory outcomes, the limited evidence available in the 

last review indicated that the strongest findings were for ambient O3 associated asthma and 

chronic obstructive pulmonary disease (COPD) respiratory hospital admissions (U.S. EPA 2013, 

6.2.7.2). Since the last review, a few additional studies have investigated cause-specific 

respiratory admissions (i.e., COPD, asthma, pneumonia) in relation to O3 exposure (MedinaRamon et al, 2006; Yang et al., 2005; Zanobetti and Schwartz, 2006; Silverman and Ito, 2010). 

Medina-Ramon et al. (2006) examined the association between short-term ambient O3

concentrations and Medicare hospital admissions for COPD among individuals ≥ 65 years of age 

for COPD in 35 cities in the U.S. for the years 1986-1999. The authors reported an increase in 

COPD admissions for lag 0-1 day in the warm season for a 30 ppb increase in 8-h max O3

concentrations. The authors found no evidence for such associations in cool season or in year 

round analyses. In a co-pollutant model with PM10, the association between O3 and COPD 

hospital admissions remained robust. In Vancouver from 1994-1998, a location with low ambient 

O3 concentrations (U.S. EPA, 2013, Table 6-26), Yang et al. (2005) reported a statistically nonsignificant increase in COPD admissions per 20 ppb increase in 24-hour average O3

concentrations. In two-pollutant models with every-day data for NO2, SO2, CO, and PM10, O3

risk estimates remained robust, though not statistically significant (U.S. EPA, 2013, Figure 6-20; 

Table 6-29). In addition, Wong et al. (2009) reported increased O3-associated COPD admissions 

during periods of increased influenza activity in Hong Kong. 

The ISA assessed a study that evaluated asthma-related hospital admissions in New York 

City (U.S. EPA, 2013, section 6.2.7.2) (Silverman and Ito, 2010). This study examined the 

association of 8-hour max O3 concentrations with severe acute asthma admissions (i.e., those 

admitted to the Intensive Care Unit [ICU]) during the warm season in the years 1999 through 
3A-3 

2006 (Silverman and Ito, 2010)). The investigators reported positive associations between O3 and 

ICU asthma admissions for the 6- to 18-year age group for a 30 ppb increase in max 8-hour 

average O3 concentrations, but little evidence of associations for the other age groups examined 

(<6 years, 19-49, 50+, and all ages). However, positive associations were observed for each 

age-stratified group and all ages for non-ICU asthma admissions, but again the strongest 

association was reported for the 6- to 18-years age group. In two-pollutant models, O3 effect 

estimates for both non-ICU and ICU hospital admissions remained robust to adjustment for 

PM2.5. In an additional analysis, using a smooth function, the authors examined whether the 

shape of the concentration-response curve for O3 and asthma hospital admissions (i.e., both 

general and ICU for all ages) is linear. When comparing the curve to a linear fit line, the authors 

found that the linear fit was a reasonable approximation of the concentration-response 

relationship between O3 and asthma hospital admissions, but the limited data density at relatively 

low O3 concentrations contributes to uncertainty in the shape of the concentration-response 

relationship at the low end of the distribution of O3 concentrations (U.S. EPA, 2013, Figure 6-

16). 

In contrast to COPD and asthma, the evidence for pneumonia-related admissions was less 

consistent. Medina-Ramon et al. (2006) examined the association between short-term ambient O3

concentrations and Medicare hospital admissions among individuals ≥ 65 years of age for 

pneumonia. The authors reported an increase in pneumonia hospital admissions in the warm 

season for a 30 ppb increase in 8-hour max O3 concentrations, with no evidence of an association 

in the cool season or year round. In two-pollutant models restricted to days for which PM10 data 

was available, the association between O3 exposure and pneumonia hospital admissions 

remained robust. In contrast, Zanobetti and Schwartz (2006) reported a decrease in pneumonia 

admissions for a 20 ppb increase in 24-hour average O3 concentrations in Boston for the average 

of lags 0 and 1 day. 

The magnitude of associations with respiratory-related hospital admissions may be 

underestimated due to behavioral modification in response to forecasted air quality (U.S. EPA, 

2013, section 4.6.6). Recent studies (Neidell and Kinney, 2010; Neidell, 2009) conducted in 

Southern California demonstrates that controlling for avoidance behavior increases O3 effect 

estimates for respiratory hospital admissions, specifically for children and older adults. This 

study shows that on days where no public alert warning of high O3 concentrations was issued, 

there was an increase in asthma hospital admissions. Although only one study has examined 

averting behavior and this study is limited to the outcome of asthma hospital admissions in one 

location and time period (i.e., Los Angeles, CA for the years 1989-1997), it does provide 

preliminary evidence indicating that some epidemiologic studies may underestimate associations 
3A-4 

between O3 and health effects by not accounting for behavioral modification when public health 

alerts are issued. 

Emergency Department Visits for All Respiratory Causes 

A large single-city study conducted in Atlanta by Tolbert et al. (2007), and subsequently 

reanalyzed by Darrow et al. (2011) using different air quality data and evaluating associations 

with different metrics, provides evidence for associations between short-term exposures to 

ambient O3 concentrations and respiratory emergency department visits. Tolbert et al. (2007) 

reported an increase in respiratory emergency department visits for a 30 ppb increase in 8-hour 

max O3 concentrations during the warm season. In copollutant models with CO, NO2, and PM10, 

limited to days in which data for all pollutants were available, associations between O3 and 

respiratory emergency department visits remained positive, but were attenuated. Darrow et al. 

(2011) reported the strongest associations with respiratory emergency department visits for 8-

hour daily max, 1-hour daily max, and day-time O3 exposure metrics (all associations positive 

and statistically significant), while positive, but statistically non-significant, associations were 

reported with 24-hour average and commuting period exposure metrics. In addition, a negative 

association was observed when using the night-time exposure metric (U.S. EPA, 2013, Figure 6-

17). The results of Darrow et al. (2011) suggest that averaging over nighttime hours may lead to 

smaller O3 effect estimates for respiratory emergency department visits due to dilution of 

relevant O3 concentrations (i.e., the higher concentrations that occur during the daytime); and 

potential negative confounding by other pollutants (e.g., CO, NO2) during the nighttime hours 

(U.S. EPA, 2013, section 6.2.7.3) 

Cause-Specific Emergency Department Visits 

In evaluating asthma emergency department visits in an all-year analysis, a Canadian 

multi-city study (Stieb et al., 2009) reported that 24-hour O3 concentrations were positively 

associated with emergency department visits for asthma at lag 1 and lag 2. Though the authors 

did not present seasonal analyses, they stated that no associations were observed with emergency 

department visits in the winter season, suggesting that the positive associations reported in the 

all-year analysis were due to the warm season (Stieb et al., 2009). In addition to asthma, the 

authors reported that O3 was positively associated with COPD emergency department visits in 

all-year analyses, but that associations with COPD visits were statistically significant only for the 

warm season (i.e., April-September). 
3A-5 

Several single-city studies have also provided evidence for positive associations between 

asthma emergency department visits and ambient O3 concentrations. Ito et al. (2007) reported 

positive and statistically significant associations with asthma emergency department visits in 

New York City during the warm season, and an inverse association in the cool season, for a 30 

ppb increase in 8-hour max O3 concentrations. In two-pollutant models with PM2.5, NO2, SO2, 

and CO, the authors found that O3 risk estimates were not substantially changed during the warm 

season (U.S. EPA, 2013, Figure 6-20; Table 6-29). 

Strickland et al. (2010) examined the association between O3 exposure and pediatric 

asthma emergency department visits (ages 5-17 years) in Atlanta using air quality data over the 

same years as Darrow et al. (2011) and Tolbert et al. (2007), but using population-weighting to 

combine daily pollutant concentrations across monitors. Strickland et al. (2010) reported an 

increase in emergency department visits for a 30 ppb increase in 8-hour max O3 concentrations 

in an all-year analysis. In seasonal analyses, stronger associations were observed during the 

warm season (i.e., May-October) than the cold season. In co-pollutant analyses that included CO, 

NO2, PM2.5 elemental carbon, or PM2.5 sulfate, Strickland et al. (2010) reported that O3 risk 

estimates were not substantially changed. The authors also examined the concentration-response 

relationship between O3 exposure and pediatric asthma emergency department visits and 

reported that positive associations with O3 persist at 8-hour ambient O3 concentrations (3-day 

average of 8-hour daily max concentrations) at least as low as 30 ppb. 

In a single-city study conducted in Seattle, WA, Mar and Koenig (2009) examined the 

association between O3 exposure and asthma emergency department visits for children (< 18) 

and adults (≥ 18). For children, positive and statistically significant associations were reported 

across multiple lags, with the strongest associations observed at lag 0 and lag 3. Ozone was also 

found to be positively associated with asthma emergency department visits for adults at all lags, 

except at lag 0. The slightly different lag times for children and adults suggest that children may 

be more immediately responsive to O3 exposures than adults (Mar and Koenig, 2009). 

In addition to the U.S. single-city studies discussed above, a single-city study conducted 

in Alberta, Canada (Villeneuve et al., 2007) provides support for the findings from Stieb et al. 

(2009), but also attempts to identify those lifestages at greatest risk for O3-associated asthma 

emergency department visits. Villeneuve et al. reported an increase in asthma emergency 

department visits in an all-year analysis across all ages with associations being stronger during 
3A-6 

the warmer months. When stratified by age, the strongest associations were observed in the 

warm season for individuals 5-14 and 15-44. These associations were not found to be 

confounded by the inclusion of aeroallergens in age-specific models. 

3.1 REFERENCES 

Biggeri, A; Baccini, M; Bellini, P; Terracini, B. (2005). Meta-analysis of the Italian studies of short-term effects of 

air pollution (MISA), 1990-1999. Int J Occup Environ Health 11: 107-122. 

Cakmak, S; Dales, RE; Judek, S. (2006b). Respiratory health effects of air pollution gases: Modification by 

education and income. Arch Environ Occup Health 61: 5-10. http://dx.doi.org/10.3200/AEOH.61.1.5-10 

Dales, RE; Cakmak, S; Doiron, MS. (2006). Gaseous air pollutants and hospitalization for respiratory disease in the 

neonatal period. Environ Health Perspect 114: 1751-1754. http://dx.doi.org/10.1289/ehp.9044 

Darrow, LA; Klein, M; Sarnat, JA; Mulholland, JA; Strickland, MJ; Sarnat, SE; Russell, AG; Tolbert, PE. (2011). 

The use of alternative pollutant metrics in time-series studies of ambient air pollution and respiratory 

emergency department visits. J Expo Sci Environ Epidemiol 21: 10-19. 

.Ito, K; Thurston, GD; Silverman, RA. (2007b). Characterization of PM2.5, gaseous pollutants, and meteorological 

interactions in the context of time-series health effects models. J Expo Sci Environ Epidemiol 17: S45-S60. 

Katsouyanni, K; Samet, JM; Anderson, HR; Atkinson, R; Le Tertre, A; Medina, S; Samoli, E; Touloumi, G; 

Burnett, RT; Krewski, D; Ramsay, T; Dominici, F; Peng, RD; Schwartz, J; Zanobetti, A. (2009). Air 

pollution and health: A European and North American approach (APHENA). (Research Report 142). 

Boston, MA: Health Effects Institute. http://pubs.healtheffects.org/view.php?id=327 

Lin, S; Bell, EM; Liu, W; Walker, RJ; Kim, NK; Hwang, SA. (2008a). Ambient ozone concentration and hospital 

admissions due to childhood respiratory diseases in New York State, 1991-2001. Environ Res 108: 42-47. 

http://dx.doi.org/10.1016/j.envres.2008.06.007 

Mar, TF; Koenig, JQ. (2009). Relationship between visits to emergency departments for asthma and ozone exposure 

in greater Seattle, Washington. Ann Allergy Asthma Immunol 103: 474-479. 

Medina-Ramon, M; Zanobetti, A; Schwartz, J. (2006). The effect of ozone and PM10 on hospital admissions for 

pneumonia and chronic obstructive pulmonary disease: A national multicity study. Am J Epidemiol 163: 

579-588. http://dx.doi.org/10.1093/aje/kwj078 

Neidell, M. (2009). Information, avoidance behavior, and health: The effect of ozone on asthma hospitalizations. 

Journal of Human Resources 44: 450-478. 

Neidell, M; Kinney, PL. (2010). Estimates of the association between ozone and asthma hospitalizations that 

account for behavioral responses to air quality information. Environ Sci Pol 13: 97-103. 

http://dx.doi.org/10.1016/j.envsci.2009.12.006 

Silverman, RA; Ito, K. (2010). Age-related association of fine particles and ozone with severe acute asthma in New 

York City. J Allergy Clin Immunol 125: 367-373. http://dx.doi.org/10.1016/j.jaci.2009.10.061 

Stieb, DM; Szyszkowicz, M; Rowe, BH; Leech, JA. (2009). Air pollution and emergency department visits for 

cardiac and respiratory conditions: A multi-city time-series analysis. Environ Health Global Access Sci 

Source 8: 25. http://dx.doi.org/10.1186/1476-069X-8-25 
3A-7 

Strickland, MJ; Darrow, LA; Klein, M; Flanders, WD; Sarnat, JA; Waller, LA; Sarnat, SE; Mulholland, JA; Tolbert, 

PE. (2010). Short-term associations between ambient air pollutants and pediatric asthma emergency 

department visits. Am J Respir Crit Care Med 182: 307-316. http://dx.doi.org/10.1164/rccm.200908-

1201OC 

Tolbert, PE; Klein, M; Peel, JL; Sarnat, SE; Sarnat, JA. (2007). Multipollutant modeling issues in a study of ambient 

air quality and emergency department visits in Atlanta. J Expo Sci Environ Epidemiol 17: S29-S35. 

http://dx.doi.org/10.1038/sj.jes.7500625 

U.S. EPA (2013). Integrated Science Assessment of Ozone and Related Photochemical Oxidants (Final Report). 

U.S. Environmental Protection Agency, Washington, DC. EPA/600/R-10/076F. Available at: 

http://www.epa.gov/ttn/naaqs/standards/ozone/s_o3_2008_isa.html 

Villeneuve, PJ; Chen, L; Rowe, BH; Coates, F. (2007). Outdoor air pollution and emergency department visits for 

asthma among children and adults: A case-crossover study in northern Alberta, Canada. Environ Health 

Global Access Sci Source 6: 40. http://dx.doi.org/10.1186/1476-069X-6-40 

Wong, CM; Yang, L; Thach, TQ; Chau, PY; Chan, KP; Thomas, GN; Lam, TH; Wong, TW; Hedley, AJ; Peiris, JS. 

(2009). Modification by influenza on health effects of air pollution in Hong Kong. Environ Health Perspect 

117: 248-253. http://dx.doi.org/10.1289/ehp.11605 

Yang, Q; Chen, Y; Krewski, D; Burnett, RT; Shi, Y; Mcgrail, KM. (2005). Effect of short-term exposure to low 

levels of gaseous pollutants on chronic obstructive pulmonary disease hospitalizations. Environ Res 99: 99-

105. http://dx.doi.org/10.1016/j.envres.2004.09.014 

Zanobetti, A; Schwartz, J. (2006). Air pollution and emergency admissions in Boston, MA. J Epidemiol Community 

Health 60: 890-895. http://dx.doi.org/10.1136/jech.2005.039834 
3B‐1

APPENDIX 3B: AMBIENT O3 CONCENTRATIONS IN LOCATIONS OF 

HEALTH STUDIES 

Annual 4th highest daily maximum O3 concentrations for all U.S. monitors operating 

during the 1975 – 2010 period were retrieved from EPA’s AQS database. These data were used 

to calculate O3 design values for the 2008 8-hour O3 NAAQS of 0.075 parts per million (ppm) 

according to 40 CFR part 50, Appendix P. Design values were calculated for each O3 monitor 

and each 3-year period between 1975-1977 and 2008-2010 whenever sufficient data were 

available. 

Ozone Design Values in Study Locations 

Ozone monitors were matched to 200 health study locations on a case-by-case basis, using 

the following guidelines: 

1) Areas defined by a Metropolitan Statistical Area (MSA) were matched with O3 monitors 

by incorporating all of the monitors located in within the MSA boundaries. 

2) Areas not represented by a MSA were matched to monitors by incorporating all of the 

monitors in the county central to location of the health study area. 

3) In some cases, EPA staff made judgment calls. For example, EPA staff matched the Los 

Angeles, CA study area to the Los Angeles-Long Beach-Santa Ana, CA MSA defined by 

Los Angeles County, CA and Orange County, CA, while the Long Beach, CA study area 

was matched to Los Angeles County, CA and the Santa Ana, CA study area was matched 

to Orange County, CA. 

 In some cases, EPA staff matched two or more study areas to the same county or MSA. 

In other cases, a study area was matched to a MSA and another study area was matched to a 

county within the same MSA. For each 3-year period, the area design value was determined by 

the monitor reporting the highest design value in the county or MSA. This has two implications 

for the design values: 

1) Design values are sensitive to changes in the monitoring network. The addition or 

discontinuation of O3 monitors in an area may cause increases or decreases in the design 

value trend. 

2) Only valid design values are reported. According to 40 CFR Part 50, Appendix P, design 

values greater than the level of the NAAQS (0.075 ppm) are always valid, while design 

values less than or equal to 0.075 ppm must have 75% annual data completeness in order 

to be valid. This may cause anomalies in the design value trend. For example, a monitor 

may report a valid design value based on as few as 12 days of data, or a monitor with less 

than 75% annual data completeness may have valid design values in some 3-year periods 

and invalid design values in others. 
3B‐2

We have identified design values for the U.S. O3 epidemiologic studies identified in Sections 

3.1.4.2 and 3.1.4.3 of the second draft Policy Assessment (see Tables 3D-1 to 3D-4). For each 

study, design values were identified for the cities evaluated and for the years over which the 

study was conducted. These design values are reported in tables A-1 to A-22 of the Wells et al, 

2012 memo “Analysis of Recent U.S. Ozone Air Quality Data to Support the O3 NAAQS 

Review and Quadratic Rollback Simulations to Support the First Draft of the Risk and Exposure 

Assessment”. 

Table 3B-1. Number of Study Cities from Multicity Epidemiologic Studies of Hospital 

Admissions and Emergency Department Visits Using Short-Term O3 Metrics with 3-Year 

Averages of Annual 4th Highest Daily Maximum 8-hour O3 Concentrations < 75 ppb1

Study Location Endpoint2

 % Increase (95% CI)3

# Study Cities < 75 

ppb over entire 

study period 

All-year

MedinaRamon et al. 

(2006)

36 U.S. cities 

COPD HA 

All year: 0.24 (-0.78, 1.21) 

Warm season: 1.63 (0.48, 

2.85) 4 

Pneumonia 

HA 

All year: 1.81 (-0.72, 4.52) 

Warm Season: 2.49 (1.57, 

3.47) 

Katsouyanni 

et al. (2009)

14 U.S. cities Respiratory 

HA 

All Year: 2.38 (0.00, 4.89) 

Warm Season: 2.14 (-0.63, 

4.97) 

2 

12 Canadian 

Cities 

Respiratory 

HA 

All year: 2.4 (0.51, 4.40) 

Warm Season: 4.1 (1.4, 

6.8) 

10 

Dales et al. 

(2006) 

11 Canadian 

cities 

Respiratory 

HA All year: 5.41 (2.88, 7.96) 7 

Stieb et al. 

(2009) 

7 Canadian 

cities 

Asthma ED All year: 3.48 (0.33, 6.76) 

5 

COPD ED 

All year: 4.03 (-0.54, 8.62) 

Warm season: 6.76 (0.11, 

13.9) 

Cakmak et al. 

(2006) 

10 Canadian 

cities 

Respiratory 

HA All year: 4.38 (2.19, 6.46) 7 

                                                            

1

 For U.S. study areas, we used EPA’s Air Quality System (AQS) (http://www.epa.gov/ttn/airs/airsaqs/) to identify 

8-hour O3 concentrations. For Canadian study areas, we used publically available air quality data from the 

Environment Canada National Air Pollution Surveillance Network (http://www.etccte.ec.gc.ca/napsdata/main/aspx). We followed the data handling protocols for calculating design values as detailed 

in 40 CFR Part 50, Appendix P. 

2

HA stands for hospital admissions; ED stands for emergency department visits. 

3

Ozone effect estimates are taken from Table 6-28 in the ISA (U.S. EPA, 2013a).  
3B‐3

Table 3B-2. Number of Study Cities from Multicity Epidemiologic Studies of Mortality 

Using Short-Term O3 Metrics with 3-Year Averages of Annual 4th Highest Daily Maximum 

8-hour O3 Concentrations < 75 ppb  

Study Location Endpoint % Increase (95% 

CI)4

# Study Cities < 75 

ppb over entire 

study period 

All-year

Schwartz 

(2005) 14 U.S. cities Non-accidental 

mortality 0.76 (0.13, 1.40) 1 

Bell et al. 

(2007) 

98 U.S. 

communities 

Non-accidental 

mortality 0.64 (0.34, 0.92) 6 

Bell and 

Dominici 

(2008) 

98 U.S. 

communities 

Non-accidental 

mortality 1.04 (0.56, 1.55) 6 

Katsouyanni 

et al. (2009) 90 U.S. cities 

Non-accidental 

mortality 3.02 (1.10, 4.89) 

6 

Respiratory 

mortality 

2.54 (-3.32, 8.79) for 

<75; 1.10 (-6.48, 9.21) 

for 75+ 

Cardiovascular 

mortality 

3.83 (-0.16, 7.95) for 

< 75; 2.30 (-1.33, 

6.04) for 75+ 

Bell et al. 

(2004) 

95 U.S. 

communities 

Non-accidental 

mortality 1.04 (0.54, 1.55) 6 

Katsouyanni 

et al. (2009) 

12 Canadian 

cities 

Non-accidental 

mortality 0.73 (0.23, 1.20) 

8 

Respiratory 

mortality 

0.13 (-1.60, 1.90); 

-0.60 (-2.70, 1.60) for 

75+ 

Cardiovascular 

mortality 

0.87 (-0.35, 2.10) for 

<75; 1.1 (0.10, 2.20) 

for 75+ 

Warm Season 

Schwartz 

(2005) 

14 U.S. 

Cities 

Non-accidental 

mortality 1.00 (0.30, 1.80) 1 

Zanobetti 

and 

Schwartz 

(2008a) 

48 U.S. cities 

Non-accidental 

mortality 1.51 (1.14, 1.87) 4 

Zanobetti 

and 

Schwartz 

(2008b) 

48 U.S. cities 

Non-accidental 

mortality 1.60 (0.84, 2.33) 

4 Respiratory 

mortality 2.51 (1.14, 3.89) 

Cardiovascular 

mortality 2.42 (1.45, 3.43) 

                                                            

4Ozone effect estimates are taken from Tables 6‐42 and 6‐53 in the ISA (U.S. EPA, 2013a).  
3B‐4

MedinaRamon and 

Schwartz 

(2008) 

48 U.S. cities 

Non-accidental 

mortality 1.96 (1.14, 2.82) 4 

Franklin and 

Schwartz 

(2008) 

18 U.S. 

communities 

Non-accidental 

mortality 1.79 (0.90, 2.68) 1 

Katsouyanni 

et al. (2009) 90 U.S. cities 

Non-accidental 

mortality 3.83 (1.90, 5.79) 

6 

Respiratory 

mortality 

4.40 (-2.10, 11.3); 

4.07 (-4.23, 13.0) for 

75+ 

Cardiovascular 

mortality 

6.78 (2.70, 11.0) for 

<75; 3.18 (-0.47, 6.95) 

for 75+ 

Bell et al. 

(2004) 

95 U.S. 

communities 

Non-accidental 

mortality 0.78 (0.26, 1.30) 6 

Katsouyanni 

et al. (2009) 

12 Canadian 

Cities 

Non-accidental 

mortality 0.42 (0.16, 0.67) 8 

 
3B‐5

Table 3B-3. Number of Study Cities from Single-City Epidemiologic Studies Using ShortTerm O3 Metrics with 3-Year Averages of Annual 4th Highest Daily Maximum 8-hour O3

Concentrations < 75 ppb  

Study Location Age Endpoint % Increase (95% 

CI)5

# Study Cities 

< 75 ppb over 

entire study 

period 

All-year 

Strickland et al. 

(2010) Atlanta Children Asthma ED 

visits 6.38 (3.19, 9.57) 0 

Warm season 

Ito et al. (2007)

New 

York City All Asthma ED visits 16.9 (10.9, 23.4) 0 

Darrow et al. 

(2011) Atlanta All Respiratory 

ED visits 2.08 (1.25, 2.91) 0 

Tolbert et al. 

(2007) Atlanta All Respiratory 

ED visits 3.90 (2.70, 5.20) 0 

Strickland et al. 

(2010) Atlanta Children Asthma ED 

visits 8.43 (4.42, 12.7) 0 

Silverman and 

Ito (2010)

New 

York City 

6 to 18 

years Asthma HA 28.2 (15.3, 41.5) 0 

All Asthma HA 12.5 (8.27, 16.7) 0 

Mar and Koenig 

(2009)

Seattle, 

WA 18+ Asthma ED 

visits 19.1 (3.00, 40.5) 1 

 

                                                            

5Ozone effect estimates are taken from Table 6‐28 in the ISA (U.S. EPA, 2013a).  
3B‐6

Table 3B-4 Number of Study Cities from Epidemiologic Studies Using Long-Term O3

Metrics with 3-Year Averages of Annual 4th Highest Daily Maximum 8-hour 

O3 Concentrations > 75, 70, 65, or 60 ppb 

Study

Number

of Cities

Study

Period

Number

(Percent) of

Cities with

Maximum

conc >75

Number

(Percent) of

Cities with

Maximum

conc >70

Number

(Percent) of

Cities with

Maximum conc

>65

Number

(Percent) of

Cities with

Maximum

conc >60

Islam et al. 2008, 2009 116 1994‐2003 11 (100%) 11 (100%) 11 (100%) 11 (100%)

Jerrett et al. 2009 947 1977‐2000 91 (97%) 92 (98%) 93 (99%) 94 (100%)

Lin et al. 2008 268 1991‐2001 24 (92%) 24 (92%) 26 (100%) 26 (100%)

Meng et al. 2010 7 1997‐2002 7 (100%) 7 (100%) 7 (100%) 7 (100%)

Moore et al. 2008 8 1980‐2000 8 (100%) 8 (100%) 8 (100%) 8 (100%)

Salam et al. 2009 119 1992‐2005 12 (100%)   12 (100%) 12 (100%) 12 (100%)

Zanobetti & Schwartz 2011 105 1985‐2006 100 (95%) 104 (99%) 104 (99%) 104 (99%)

 

                                                            

6 Study authors included 12 cities in their analyses, air quality data that met completeness criteria described above

were available for 11 cities 7 Study authors included 96 cities in their analyses, air quality data that met completeness criteria described above  

were available for 94 cities 8 Study authors included 27 cities in their analyses, air quality data that met completeness criteria described above

were available for 26 cities 9 Study authors included 12 cities in their analyses, air quality data that met completeness criteria described above

were available for 11 cities
3B‐7

Relationship between average 24-hour and highest 8-hour O3 concentrations for cities 

analyzed by Bell et al. (2006) 

Bell et al. (2006) reported associations between mortality and 24-hour average O3

concentrations (i.e., averaged across monitors in cities with multiple monitors) in a multi-city 

study of 98 U.S. cities. Positive associations persisted in a series of analyses that restricted O3

concentrations to those below various cut points (cut points ranged from 5 to 60 ppb in 5 ppb 

increments). To facilitate consideration of these cut point analyses for the second draft of the O3

Policy Assessment, so as to match the form and averaging time of the existing primary standard, 

we evaluated the relationship between 24-hour average O3 concentrations, averaged across 

monitors in cities with multiple monitors, and the highest 8-hour daily maximum O3

concentrations among the individual monitors in each city. 

EPA staff retrieved daily 24-hour average and 8-hour maximum O3 concentrations 

reported to EPA by monitors in the 98 study areas defined in Bell et al. (2006) during the 1987-

2012 period from EPA’s Air Quality System (AQS) database. Next, EPA staff obtained the 

study area boundaries from the published study (Bell et al., 2006) and used them to determine 

which O3 monitoring sites were associated with each study area. The 24-hour average O3

concentrations were averaged spatially across all available monitors within each study area on 

each day where monitoring data were collected. Next, days where the area-wide 24-hour average 

concentration (i.e., averaged spatially across monitors in areas with multiple monitors) was 

greater than 60 ppb were removed from the data. Based on the data remaining (i.e., with 24-hour 

average concentrations of 60 ppb or below), the annual 4th highest 8-hour daily maximum 

concentrations were identified for each study area and for each year from 1987-2012 (Table 3D3). This process was repeated by further removing days with area-wide 24-hour average 

concentrations greater than 55 ppb, 50 ppb, etc., down to 5 ppb, and re-calculating the same 

statistics after each removal. The resulting dataset consisted of the annual 4th highest 8-hour 

daily maximum concentrations for all study areas. 

 
3B‐8

Table 3B-5 Number of Study Cities with 4th Highest 8-hour Daily Maximum 

Concentrations Greater Than the Level of the Current Standard and 

Potential Alternative Standards For Various Cut-Point Analyses Presented 

in Bell et al. (2006)10

 

                                                            

10 Study authors included 98 cities in their analyses, air quality data only available for 95

20 25 30 35 40 45 50 55 60 All

Number (%) of Cities

with 4th highest >75

(any year; 1987‐2000)

0 (0%) 0 (0%) 12 (12%) 52 (53%) 77 (79%) 88 (90%) 93 (95%) 94 (96%) 94 (96%) 94 (96%)

Number (%) of Cities

with 4th highest >70

(any year; 1987‐2000)

0 (0%) 3 (3%) 31 (32%) 77 (79%) 86 (88%) 93 (95%) 94 (96%) 94 (96%) 95 (97%) 95 (97%)

Number (%) of Cities

with 4th highest >65

(any year; 1987‐2000)

0 (0%) 10 (10%) 58 (59%) 84 (86%) 93 (95%) 94 (96%) 94 (96%) 94 (96%) 94 (96%) 94 (96%)

Number (%) of Cities

with 4th highest >60

(any year; 1987‐2000)

1 (1%) 36 (37%) 74 (76%) 93 (95%) 96 (8%) 97 (99%) 97 (99%) 97 (99%) 97 (99%) 97 (99%)

Cut‐point for 2‐day moving average across monitors and cities (24‐h avg)
3B‐9

Relationship between average and highest 8-hour daily maximum O3 concentrations for 

New York City, as analyzed by Silverman and Ito (2010) 

EPA staff retrieved daily maximum 8-hour O3 concentrations for the 13 monitors in the 

New York City area used in the Silverman and Ito (2010) study for April-August of 1999-2006 

from the AQS database. Next, EPA staff spatially averaged these concentrations across monitors 

for each day during this period, and then paired them with the highest 8-hour daily maximum 

value reported across the 13 monitors on each day. 

Next, the range of observed average daily maximum 8-hour concentrations was broken 

into 5 ppb increments. The number of days where the area-wide average daily maximum 8-hour 

concentration fell within the increment and the number of days where one or more monitored 8-

hour daily maximum values were greater than 75, 70, 65 and 60 ppb were recorded for each 5 

ppb increment. These numbers are summarized in Table 3D-4. 

Table 3B-6 Summary statistics for Observed O3 Concentrations in the New York City 

Area, April – August 1999 – 2006 

 

11 to 20

(62 days)

21 to 25

(92 days)

26 to 30

(178 days)

31 to 35

(206 days)

36 to 40

(236 days)

41 to 45

(196 days)

46 to 50

(153 days)

51 to 55

(111 days)

56 to 60

(71 days)

Days > 75 ppb 0 0 1 0 1 2 9 15 20

Days > 70 ppb 0 0 1 4 1 12 17 23 30

Days > 65 ppb 0 0 1 6 5 18 37 42 45

Days > 60 ppb 0 0 2 7 12 39 67 61 53

2‐day moving average across monitors (ppb)
3B‐10

Relationship between average and highest 8-hour daily maximum O3 concentrations for 

Atlanta, as analyzed by Strickland et al. (2010) 

For our assessment of the Strickland et al. (2010) study, based in the Atlanta metropolitan 

area, we retrieved 8-hour daily maximum concentration data for 4 of the 5 monitors used in the 

study during the study period (May-October, 1993-2004) from the AQS database. The 5th

monitor was a part of the Southeastern Aerosol Research and Characterization (SEARCH) 

network, which does not report data to EPA. EPA staff calculated the area-wide average of the 

8-hour daily maximum concentrations for each day, and compared to population-weighted 

average concentrations obtained from the author. The correlation between the arithmetic average 

values and the population-weighted average values was very high (R = 0.985), thus EPA staff 

deemed the arithmetic average to be a suitable surrogate for the population-weighted average 

used in the study. Finally, 3-day moving averages were calculated from the daily area-wide 

average values (matching the air quality metric used in the study), and paired with the highest 

monitored 8-hour daily maximum value occurring during each 3-day period. 

Next, the range of observed average daily maximum 8-hour concentrations was broken 

into 5 ppb increments. The number of days where the area-wide average daily maximum 8-hour 

concentration fell within the increment and the number of days where one or more monitored 8-

hour daily maximum values were greater than 75, 70, 65 and 60 ppb were recorded for each 5 

ppb increment. These numbers are summarized in Table 3D-5. 

Table 3B-7 Summary statistics for Observed O3 Concentrations in the Atlanta Area, 

April – August 1999 – 2006 

 

26‐30      

(75 days)

31‐35      

(144 days)

36‐40    

(165 days)

41‐45    

(210 days)

46‐50     

(235 days)

51‐55     

(244 days)

56‐60     

(272 days)

61‐65     

(234 days)

66‐70      

(169 days)

71 to 75

(124 days)

76 to 80

(106 days)

Days > 75 0 0 2 2 10 24 53 80 89 87 87

Days > 70 0 0 6 6 20 49 81 111 107 96 95

Days > 65 1 0 8 19 38 75 118 147 133 106 100

Days > 60 1 2 15 33 68 115 152 173 147 116 102

3‐day moving average across monitors (ppb)
3B‐11

Relationship between annual and highest 1-hour daily maximum O3 concentrations for 12 

study areas, as analyzed by Jerrett et al. (2009) 

The Jerrett et al. (2009) study used a long-term metric based on seasonal averages of 1-

hour daily maximum O3 concentrations to evaluate associations between respiratory mortality 

and long-term or repeated exposures to O3. Authors divided study cities into quartiles based on 

these seasonal averages of 1-hour daily O3 concentrations. Using AQS, we identified the 3-year 

averages of annual 4th highest daily maximum 8-hour O3 concentrations in study cities during the 

study period. Table 3D-6 presents the means and maximums of these concentrations over the 

study period. 

In addition, for the 12 urban case study areas included in the epidemiology-based risk 

assessment of the 2nd draft of the Health REA we identified the seasonal averages of 1-hour daily 

maximum concentrations (i.e., the O3 metric evaluated by Jerrett et al., 2009) for air quality 

adjusted to the current and alternative standards. Specifically, for adjusted air quality “quarterly” 

averages of 1-hour concentrations for April-June and July-August were calculated for each area 

and year. The quarterly values were considered to be valid if valid daily maximum 1-hour 

values were available for at least 75% of the days in the quarter. The two quarterly values were 

then averaged, as was done by Jerrett et al. (2009) to generate the long-term metric used in the 

study. This process was repeated for the various model-based adjustment scenarios in each of 

the 12 study areas. Summary statistics based on this seasonal average of daily O3 concentrations 

are presented in Table 3D-7 for recent air quality and for air quality adjusted to just meet the 

current and alternative standards. 

 
3B‐12

Table 3B-8 Three-Year Averages of Annual 4th Highest Daily Maximum 8-hour O3

Concentrations in 9411 Study Areas Examined in Jerrett et al. (2009) 

City Mean over

study

period

Max  over

study

period

Cities in the lowest quartile of average exposure12

Charleston, WV 81 99

Chicago, IL 103 114

Colorado Springs, CO 62 66

Corpus Christi, TX 82 89

Detroit, MI 95 103

Flint, MI 83 91

Ft. Lauderdale, FL 74 79

Kansas City, MO 87 97

Lansing, MI 81 90

Madison, WI 82 102

Minneapolis, MN 74 80

New Orleans, LA 86 99

Orlando, FL 79 82

Portland, OR 81 91

Providence, RI 110 124

Salinas, CA 68 74

San Antonio, TX 85 92

San Francisco, CA 88 96

San Jose, CA 91 103

Seattle, WA 78 88

Tacoma, WA 78 88

Vallejo, CA 74 82

Wichita, KS 75 81

Cities in the highest three

quartiles of average

exposure13

Charleston, SC 79 90

Charlotte, NC 97 112

Chattanooga, TN 90 97

Cincinnati, OH 101 119

Cleveland, OH 98 108

Columbia, SC 85 109

Columbus, OH 93 103

Dallas/Ft Worth, TX 106 118

Dayton, OH 95 122

Denver, CO 83 91

                                                            

11 Jerrett et al. (2009) examined 96 MSAs; this analysis included the 94 cities that met data completeness criteria

described above, after linking monitors to MSAs (see lines 10‐28, above). 12 Based on visual inspection of Figure 1 in Jerrett et al. (2009) 13 Based on visual inspection of Figure 1 in Jerrett et al. (2009)
3B‐13

El Paso, TX 85 96

Evansville, IN 93 100

Fresno, CA 112 123

Gary, IN 91 105

Greely, CO 69 75

Greensboro, NC 89 100

Greenville, SC 86 94

Harrisburg, PA 94 103

Houston, TX 121 140

Huntington, WV 94 103

Indianapolis, IN 93 103

Jackson, MS 79 98

Jacksonville, FL 81 87

Jersey City, NJ 106 118

Johnstown, PA 90 107

Kenosha, WI 101 114

Knoxville, TN 91 97

Lancaster, PA 94 101

Las Vegas, NV 80 85

Lexington, KY 88 99

Little Rock, AR 86 107

Los Angeles, CA 193 248

Memphis, TN 94 103

Milwaukee, WI 103 117

Nashville, TN 94 106

Nassau, NY NA14 NA

New Haven, CT 116 136

New York City, NY 118 129

Newark, NJ 90 105

Norfolk, VA 91 101

Oklahoma City, OK 86 93

Philadelphia, PA 117 136

Phoenix, AZ 86 96

Pittsburgh, PA 101 123

Portland, ME 106 117

Portsmouth, NH 92 104

Racine, WI 102 124

Raleigh, NC 90 104

Reading, PA 99 114

Richmond, VA 94 104

Riverside, CA 196 245

Roanoke, VA 83 95

                                                            

14 Air quality data did not meet completeness criteria described above
3B‐14

Rochester, NY 89 99

Sacramento, CA 110 118

San Diego, CA 121 141

Shreveport, LA 83 88

South Bend, IN 90 102

Springfield, MA 102 115

St Louis, MO 105 122

Steubenville, OH 82 99

Syracuse, NY 85 96

Tampa, FL 85 91

Toledo, OH 93 108

Trenton, NJ 112 124

Tucson, AZ 76 82

Ventura, CA 118 132

Washington, DC 105 116

Wilmington, DE 103 116

Worcester, MA 92 102

York, PA 95 107

Youngstown, OH 93 103

 
3B‐15

Table 3B-9 Long-Term O3 Concentrations in 12 Urban Case Study Areas (Using the O3

Metric Evaluated by Jerrett et al., 2009) for Recent Air Quality and Air 

Quality Adjusted to Meet Standard Levels of 75, 70, 65, and 60 ppb 

 

Air Quality

Adjusted to:

2006                 

(Adj Yrs 2006‐2008)

2007                

(Adj Yrs 2006‐2008)

2008                

(Adj Yrs 2008‐2010)

2009                  

(Adj Yrs 2008‐2010)

2010                 

(Adj Yrs 2008‐2010)

Recent 65 63 57 50 56

75 53 52 53 47 52

70 50 49 49 44 49

65 47 46 46 42 46

60 45 44 44 40 44

Recent 60 59 57 52 60

75 54 54 53 49 55

70 52 51 51 48 53

65 49 49 48 46 50

60 46 46 46 44 48

Recent 49 50 46 45 49

75 48 49 49 45 48

70 46 47 48 44 48

65 44 45 46 43 46

60 43 43 44 41 44

Recent 51 52 53 49 54

75 49 50 51 47 51

70 47 48 48 45 48

65 45 45 45 43 45

60 41 41 41 40 42

Recent 63 63 63 58 60

75 62 61 63 58 60

70 60 59 62 58 58

65 58 58 59 56 55

60 53 53 53 51 50

Recent 50 54 51 48 52

75 50 52 N/A N/A N/A

70 48 50 51 49 52

65 47 49 49 47 50

60 45 46 46 45 47

Recent 53 48 47 47 46

75 48 46 47 48 46

70 47 45 46 47 46

65 46 44 45 46 45

60 45 43 43 44 44

Recent 65 61 64 62 57

75 58 59 60 60 58

70 55 56 57 58 56

65 52 53 54 54 53

60 N/A N/A N/A N/A N/A

Recent 53 54 55 48 55

75 47 47 51 47 51

70 N/A N/A N/A N/A N/A

65 N/A N/A N/A N/A N/A

60 N/A N/A N/A N/A N/A

Recent 56 59 57 51 58

75 51 52 54 49 54

70 49 50 51 47 52

65 47 48 49 45 49

60 45 46 47 43 47

Recent 66 59 65 61 55

75 55 50 54 51 48

70 52 48 51 49 46

65 50 46 49 47 44

60 47 44 46 44 42

Recent 58 58 52 51 55

75 53 53 51 50 54

70 50 51 50 48 52

65 47 48 48 46 49

60 44 45 45 43 46

Houston

Philadelphia

Sacramento

Saint Louis

Atlanta

Baltimore

Boston

Cleveland

Denver

Detroit

Los Angeles

New York City
3B‐16

REFERENCES 

Bell, ML; Dominici, F. (2008). Effect modification by community characteristics on the short-term effects of ozone 

exposure and mortality in 98 US communities. Am J Epidemiol 167: 986-997. 

http://dx.doi.org/10.1093/aje/kwm396 

Bell, ML; Kim, JY; Dominici, F. (2007). Potential confounding of particulate matter on the short-term association 

between ozone and mortality in multisite time-series studies. Environ Health Perspect 115: 1591-1595. 

http://dx.doi.org/10.1289/ehp.10108 

Bell, ML; Peng, RD; Dominici, F. (2006). The exposure-response curve for ozone and risk of mortality and the 

adequacy of current ozone regulations. Environ Health Perspect 114: 532-536. 

Cakmak, S; Dales, RE; Judek, S. (2006a). Do gender, education, and income modify the effect of air pollution gases 

on cardiac disease? J Occup Environ Med 48: 89-94. 

http://dx.doi.org/10.1097/01.jom.0000184878.11956.4b 

Cakmak, S; Dales, RE; Judek, S. (2006b). Respiratory health effects of air pollution gases: Modification by 

education and income. Arch Environ Occup Health 61: 5-10. http://dx.doi.org/10.3200/AEOH.61.1.5-10 

Dales, RE; Cakmak, S; Doiron, MS. (2006). Gaseous air pollutants and hospitalization for respiratory disease in the 

neonatal period. Environ Health Perspect 114: 1751-1754. http://dx.doi.org/10.1289/ehp.9044 

Franklin, M; Schwartz, J. (2008). The impact of secondary particles on the association between ambient ozone and 

mortality. Environ Health Perspect 116: 453-458. http://dx.doi.org/10.1289/ehp.10777 

Islam, T; Berhane, K; McConnell, R; Gauderman, WJ; Avol, E; Peters, JM; Gilliland, FD. (2009). Glutathione-Stransferase (GST) P1, GSTM1, exercise, ozone and asthma incidence in school children. Thorax 64: 197-

202. http://dx.doi.org/10.1136/thx.2008.099366 

Islam, T; McConnell, R; Gauderman, WJ; Avol, E; Peters, JM; Gilliland, FD. (2008). Ozone, oxidant defense genes 

and risk of asthma during adolescence. Am J Respir Crit Care Med 177: 388-395. 

http://dx.doi.org/10.1164/rccm.200706-863OC 

Jerrett, M; Burnett, RT; Pope, CA, III; Ito, K; Thurston, G; Krewski, D; Shi, Y; Calle, E; Thun, M. (2009). Longterm ozone exposure and mortality. N Engl J Med 360: 1085-1095. 

http://dx.doi.org/10.1056/NEJMoa0803894 

Katsouyanni, K; Samet, JM; Anderson, HR; Atkinson, R; Le Tertre, A; Medina, S; Samoli, E; Touloumi, G; 

Burnett, RT; Krewski, D; Ramsay, T; Dominici, F; Peng, RD; Schwartz, J; Zanobetti, A. (2009). Air 

pollution and health: A European and North American approach (APHENA). (Research Report 142). 

Boston, MA: Health Effects Institute. http://pubs.healtheffects.org/view.php?id=327 

Lin, S; Liu, X; Le, LH; Hwang, SA. (2008b). Chronic exposure to ambient ozone and asthma hospital admissions 

among children. Environ Health Perspect 116: 1725-1730. http://dx.doi.org/10.1289/ehp.11184 

Medina-Ramon, M; Zanobetti, A; Schwartz, J. (2006). The effect of ozone and PM10 on hospital admissions for 

pneumonia and chronic obstructive pulmonary disease: A national multicity study. Am J Epidemiol 163: 

579-588. http://dx.doi.org/10.1093/aje/kwj078 

Meng, YY; Rull, RP; Wilhelm, M; Lombardi, C; Balmes, J; Ritz, B. (2010). Outdoor air pollution and uncontrolled 

asthma in the San Joaquin Valley, California. J Epidemiol Community Health 64: 142-147. 

http://dx.doi.org/10.1136/jech.2008.083576 
3B‐17

Moore, K; Neugebauer, R; Lurmann, F; Hall, J; Brajer, V; Alcorn, S; Tager, I. (2008). Ambient ozone 

concentrations cause increased hospitalizations for asthma in children: An 18-year study in Southern 

California. Environ Health Perspect 116: 1063-1070. http://dx.doi.org/10.1289/ehp.10497

Salam, MT; Islam, T; Gauderman, WJ; Gilliland, FD. (2009). Roles of arginase variants, atopy, and ozone in 

childhood asthma. J Allergy Clin Immunol 123: 596-602. http://dx.doi.org/10.1016/j.jaci.2008.12.020 

Schwartz, J. (2005a). How sensitive is the association between ozone and daily deaths to control for temperature? 

Am J Respir Crit Care Med 171: 627-631. 

Schwartz, J. (2005b). Who is sensitive to extremes of temperature? A case-only analysis. Epidemiology 16: 67-72. 

http://dx.doi.org/10.1097/01.ede.0000147114.25957.71 

Stieb, DM; Szyszkowicz, M; Rowe, BH; Leech, JA. (2009). Air pollution and emergency department visits for 

cardiac and respiratory conditions: A multi-city time-series analysis. Environ Health Global Access Sci 

Source 8: 25. http://dx.doi.org/10.1186/1476-069X-8-25 

Zanobetti, A; Schwartz, J. (2011). Ozone and survival in four cohorts with potentially predisposing diseases. Am J 

Respir Crit Care Med 184: 836-841. http://dx.doi.org/10.1164/rccm.201102-0227OC 

Zanobetti, A; Schwartz, J. (2008a). Is there adaptation in the ozone mortality relationship: A multi-city casecrossover analysis. Environ Health 7: 22. http://dx.doi.org/10.1186/1476-069X-7-22 

Zanobetti, A; Schwartz, J. (2008b). Mortality displacement in the association of ozone with mortality: An analysis 

of 48 cities in the United States. Am J Respir Crit Care Med 177: 184-189. 

http://dx.doi.org/10.1164/rccm.200706-823OC 
5A-1 

Appendix 5A 

Ozone-Sensitive Plant SpeciesA Used by Some Tribes* 

*(Based on Feedback from 3 Tribes) 

Common Name 

(other common names)

Scientific Name Confirmed bioindicator 

species 

Red alder (Oregon alder, Western alder) Alnus rubra Y 

Speckled alder (Tag alder, Gray alder, Hoary 

alder) Alnus rugosa (Alnus incana) Y 

Groundnut (Wild bean, American potato bean) Apios americana Y 

Spreading Dogbane (Common dogbane) Apocynum androsamifolium Y 

Common milkweed Asclepias syriaca Y 

New England Aster Aster novae-angliae 

Symphyotrichum novae-angliae 

Green ash Fraxinus pennsylvanica 

Twinberry Lonicera involucrate Y 

Bee-balm Monarda didyma 

Virginia creeper Parthenocissus quinquefolia Y 

Jack pine Pinus banksiana Y 

Lodgepole pine Pinus contorta 

White pine Pinus strobus 

Black poplar (Balsam poplar) Populus balsamifera 

trichocarpa 

Quaking aspen (Trembling aspen) Populus tremuloides Y 

Black cherry Prunus serotina Y 

Choke cherry Prunus virginiana 

Douglas fir Pseudotsuga menziesii 

Allegheny blackberry (Common blackberry) Rubus allegheniensis Y 

Thimbleberry Rubus parviflorus Y 

Cutleaf coneflower (Coneflower, Golden glow) Rudbeckia laciniata Y 

Pussy willow Salix discolor 

Shinning willow Salix lucida 

American elder (White elder) Sambucus canadensis Y 

Red elderberry Sambucus racemosa Y 

Sassafras Sassafras albidum 

Goldenrod Solidago altissima 

Huckleberry Vaccinium membranaceum Y 

Wild grape Vitis spp. 

European wine grape Vitis vinifera Y 

ASpecies included in this list are identified in one or more of the following sources: 

1)SP 2007 (www.2.nature.nps.gov/air/Pubs/pdf/flag/NPSozonesensppFLAG06.pdf) 

2) NPS O3 Bioindicators 2006 (www.nature.nps.gov/air/Pubs/bioindicators/index.cfm) 

3) Kline et al., 2008; 4) Davis, 2007/ 2009; 5) Flagler, et al., eds., 1998 

6) USDA FS FHM/FIA: Ozone Bioindicator Sampling and Estimation 

(www.nrs.fs.fed.us/fia/topics/ozone/pubs/pdfs/ozone%20estimation%20document.pdf) and 

Ozone Injury in West Coast Forests: 6 Years of Monitoring (2007).
5A-2 

This page left intentionally blank.
5B-1 

APPENDIX 5B: CLASS I AREAS BELOW CURRENT STANDARD AND 

ABOVE 15 PPM-HRS 

This appendix identifies Class I areas that might have W126 index values above 15 ppmhrs allowed by the current standard based on an analysis of recent O3 monitoring data. Table 5B1 provides all monitoring sites from 1998-2002 that were at or below 75 ppb (3-year average of 

4th highest maximum 8-hour average), at or above 15 ppm-hrs (3-year average), and located in 

counties with Class I areas. For each year that met these 3-year requirements, we also provide 

the maximum annual 8-hour O3 concentration (in ppb) and W126 index value (in ppm-hrs).
5B-2 

Table 5B-1 Examples of Counties Containing Class I Areas where Recent 3-Year O3 Concentrations were Below 

W126 Index Values were Above 15 ppm-hrs 

Monitor ID 

# 

Years 

(3-year 

average) 

Year 

(annual) 

Max 8-hour 

(ppb) 

(3-year 

average) 

Max 8-hour

(ppb) 

(annual) 

W126 

(3-year 

average) 

W126 

(annual) Monitor Site Name State County 

400380011 1998-2000 1998 70 67 15.88 14.71 Chiricahua National Monument AZ Cochise 400380011 1998-2000 1999 70 72 15.88 16.57 Chiricahua National Monument AZ Cochise 400380011 1998-2000 2000 70 71 15.88 16.36 Chiricahua National Monument AZ Cochise 400380011 2002-2004 2002 71 74 15.70 14.45 Chiricahua National Monument AZ Cochise 400380011 2002-2004 2003 71 71 15.70 18.07 Chiricahua National Monument AZ Cochise 400380011 2002-2004 2004 71 70 15.70 14.57 Chiricahua National Monument AZ Cochise 400380011 2003-2005 2003 71 71 16.64 18.07 Chiricahua National Monument AZ Cochise 400380011 2003-2005 2004 71 70 16.64 14.57 Chiricahua National Monument AZ Cochise 400380011 2003-2005 2005 71 72 16.64 17.29 Chiricahua National Monument AZ Cochise 400380011 2004-2006 2004 72 70 16.56 14.57 Chiricahua National Monument AZ Cochise 400380011 2004-2006 2005 72 72 16.56 17.29 Chiricahua National Monument AZ Cochise 400380011 2004-2006 2006 72 74 16.56 17.81 Chiricahua National Monument AZ Cochise 400380011 2005-2007 2005 71 72 16.36 17.29 Chiricahua National Monument AZ Cochise 400380011 2005-2007 2006 71 74 16.36 17.81 Chiricahua National Monument AZ Cochise 400380011 2005-2007 2007 71 67 16.36 13.98 Chiricahua National Monument AZ Cochise 400380011 2006-2008 2006 69 74 16.37 17.81 Chiricahua National Monument AZ Cochise 400380011 2006-2008 2007 69 67 16.37 13.98 Chiricahua National Monument AZ Cochise 400380011 2006-2008 2008 69 68 16.37 17.32 Chiricahua National Monument AZ Cochise 400380011 2010-2012 2010 73 71 18.06 13.21 Chiricahua National Monument AZ Cochise 400380011 2010-2012 2011 73 75 18.06 19.33 Chiricahua National Monument AZ Cochise 400380011 2010-2012 2012 73 74 18.06 21.65 Chiricahua National Monument AZ Cochise 400510081 2008-2010 2008 69 74 15.61 22.20 Flagstaff Middle School AZ Coconino 400510081 2008-2010 2009 69 66 15.61 11.38 Flagstaff Middle School AZ Coconino 400580011 2008-2010 2010 69 69 15.61 14.89 Flagstaff Middle School AZ Coconino 400510081 2006-2008 2008 70 74 19.29 22.20 Grand Canyon National Park, The Abyss AZ Coconino 400510081 2007-2009 2008 68 74 15.38 22.20 Grand Canyon National Park, The Abyss AZ Coconino 400510081 2007-2009 2009 68 66 15.38 11.38 Grand Canyon National Park, The Abyss AZ Coconino 400580011 1998-2000 1998 73 72 18.74 18.23 Grand Canyon National Park, The Abyss AZ Coconino 400580011 1998-2000 1999 73 76 18.74 21.27 Grand Canyon National Park, The Abyss AZ Coconino 400580011 1998-2000 2000 73 71 18.74 16.74 Grand Canyon National Park, The Abyss AZ Coconino 400580011 1999-2001 1999 72 76 17.64 21.27 Grand Canyon National Park, The Abyss AZ Coconino 400580011 1999-2001 2000 72 71 17.64 16.74 Grand Canyon National Park, The Abyss AZ Coconino 
5B-3 

Monitor ID 

# 

Years 

(3-year 

average) 

Year 

(annual) 

Max 8-hour 

(ppb) 

(3-year 

average) 

Max 8-hour

(ppb) 

(annual) 

W126 

(3-year 

average) 

W126 

(annual) Monitor Site Name State County 

400580011 1999-2001 2001 72 70 17.64 14.91 Grand Canyon National Park, The Abyss AZ Coconino 400580011 2000-2002 2000 73 71 19.47 16.74 Grand Canyon National Park, The Abyss AZ Coconino 400580011 2000-2002 2001 73 70 19.47 14.91 Grand Canyon National Park, The Abyss AZ Coconino 400580011 2000-2002 2002 73 79 19.47 26.78 Grand Canyon National Park, The Abyss AZ Coconino 400580011 2001-2003 2001 74 70 21.79 14.91 Grand Canyon National Park, The Abyss AZ Coconino 400580011 2001-2003 2002 74 79 21.79 26.78 Grand Canyon National Park, The Abyss AZ Coconino 400580011 2001-2003 2003 74 73 21.79 23.70 Grand Canyon National Park, The Abyss AZ Coconino 400580011 2002-2004 2002 74 79 22.29 26.78 Grand Canyon National Park, The Abyss AZ Coconino 400580011 2002-2004 2003 74 73 22.29 23.70 Grand Canyon National Park, The Abyss AZ Coconino 400580011 2002-2004 2004 74 72 22.29 16.41 Grand Canyon National Park, The Abyss AZ Coconino 400580011 2003-2005 2003 74 73 19.98 23.70 Grand Canyon National Park, The Abyss AZ Coconino 400580011 2003-2005 2004 74 72 19.98 16.41 Grand Canyon National Park, The Abyss AZ Coconino 400580011 2003-2005 2005 74 79 19.98 19.84 Grand Canyon National Park, The Abyss AZ Coconino 400580011 2004-2006 2004 73 72 19.39 16.41 Grand Canyon National Park, The Abyss AZ Coconino 400580011 2004-2006 2005 73 79 19.39 19.84 Grand Canyon National Park, The Abyss AZ Coconino 400580011 2004-2006 2006 73 70 19.39 21.92 Grand Canyon National Park, The Abyss AZ Coconino 400580011 2005-2007 2005 72 79 20.24 19.84 Grand Canyon National Park, The Abyss AZ Coconino 400580011 2005-2007 2006 72 70 20.24 21.92 Grand Canyon National Park, The Abyss AZ Coconino 400580011 2005-2007 2007 72 69 20.24 18.95 Grand Canyon National Park, The Abyss AZ Coconino 400580011 2006-2008 2006 70 70 19.29 21.92 Grand Canyon National Park, The Abyss AZ Coconino 400580011 2006-2008 2007 70 69 19.29 18.95 Grand Canyon National Park, The Abyss AZ Coconino 400580011 2007-2009 2007 68 69 15.38 18.95 Grand Canyon National Park, The Abyss AZ Coconino 400580011 2010-2012 2010 72 69 17.90 14.89 Grand Canyon National Park, The Abyss AZ Coconino 400580011 2010-2012 2011 72 74 17.90 18.45 Grand Canyon National Park, The Abyss AZ Coconino 400580011 2010-2012 2012 72 73 17.90 20.34 Grand Canyon National Park, The Abyss AZ Coconino 400700101 2007-2009 2007 75 76 22.45 24.95 Tonto NM AZ Gila 400700101 2007-2009 2008 75 78 22.45 27.52 Tonto NM AZ Gila 400700101 2007-2009 2009 75 72 22.45 14.89 Tonto NM AZ Gila 400700101 2008-2010 2008 73 78 20.29 27.52 Tonto NM AZ Gila 400700101 2008-2010 2009 73 72 20.29 14.89 Tonto NM AZ Gila 400700101 2008-2010 2010 73 70 20.29 18.45 Tonto NM AZ Gila 400700101 2009-2011 2009 72 72 17.90 14.89 Tonto NM AZ Gila 400700101 2009-2011 2010 72 70 17.90 18.45 Tonto NM AZ Gila 400700101 2009-2011 2011 72 76 17.90 20.35 Tonto NM AZ Gila 400700101 2010-2012 2010 74 70 21.41 18.45 Tonto NM AZ Gila 
5B-4 

Monitor ID 

# 

Years 

(3-year 

average) 

Year 

(annual) 

Max 8-hour 

(ppb) 

(3-year 

average) 

Max 8-hour

(ppb) 

(annual) 

W126 

(3-year 

average) 

W126 

(annual) Monitor Site Name State County 

400700101 2010-2012 2011 74 76 21.41 20.35 Tonto NM AZ Gila 400700101 2010-2012 2012 74 78 21.41 25.44 Tonto NM AZ Gila 401320051 2007-2009 2009 75 70 22.48 14.51 Rio Verde AZ Maricopa 401340081 2007-2009 2008 75 78 22.48 27.49 Rio Verde AZ Maricopa 401397061 2007-2009 2007 75 79 22.48 28.65 Rio Verde AZ Maricopa 401701191 2010-2012 2010 70 68 15.79 12.96 Petrified Forest National Park, South Entrance AZ Navajo 401701191 2010-2012 2011 70 69 15.79 15.16 Petrified Forest National Park, South Entrance AZ Navajo 401701191 2010-2012 2012 70 73 15.79 19.26 Petrified Forest National Park, South Entrance AZ Navajo 401900211 1998-2000 1998 73 77 15.55 18.60 22nd & Craycroft AZ Pima 401910181 1998-2000 1999 73 73 15.55 16.53 22nd & Craycroft AZ Pima 401910281 1998-2000 2000 73 77 15.55 15.52 22nd & Craycroft AZ Pima 401900211 2001-2003 2002 73 77 15.53 16.01 Saguaro Park AZ Pima 401900211 2001-2003 2003 73 78 15.53 23.14 Saguaro Park AZ Pima 401900211 2006-2008 2007 74 73 18.98 17.24 Saguaro Park AZ Pima 401900211 2006-2008 2008 74 74 18.98 20.01 Saguaro Park AZ Pima 401900211 2007-2009 2007 71 73 16.10 17.24 Saguaro Park AZ Pima 401900211 2007-2009 2008 71 74 16.10 20.01 Saguaro Park AZ Pima 401900211 2007-2009 2009 71 67 16.10 11.04 Saguaro Park AZ Pima 401900211 2008-2010 2008 69 74 15.47 20.01 Saguaro Park AZ Pima 401900211 2008-2010 2009 69 67 15.47 11.04 Saguaro Park AZ Pima 401900211 2008-2010 2010 69 68 15.47 15.36 Saguaro Park AZ Pima 401900211 2010-2012 2010 71 68 16.84 15.36 Saguaro Park AZ Pima 401900211 2010-2012 2011 71 75 16.84 17.36 Saguaro Park AZ Pima 401900211 2010-2012 2012 71 71 16.84 17.79 Saguaro Park AZ Pima 401910111 2001-2003 2001 73 69 15.53 12.73 Saguaro Park AZ Pima 401910181 2006-2008 2006 74 76 18.98 21.54 Saguaro Park AZ Pima 402130011 2007-2009 2007 75 77 22.52 24.59 Queen Valley AZ Pinal 402180011 2007-2009 2008 75 80 22.52 29.02 Queen Valley AZ Pinal 402180011 2007-2009 2009 75 70 22.52 14.81 Queen Valley AZ Pinal 402180011 2008-2010 2008 74 80 20.87 29.02 Queen Valley AZ Pinal 402180011 2008-2010 2009 74 70 20.87 14.81 Queen Valley AZ Pinal 402180011 2008-2010 2010 74 72 20.87 18.79 Queen Valley AZ Pinal 402180011 2009-2011 2009 73 70 18.75 14.81 Queen Valley AZ Pinal 402180011 2009-2011 2010 73 72 18.75 18.79 Queen Valley AZ Pinal 402180011 2009-2011 2011 73 78 18.75 22.66 Queen Valley AZ Pinal 
5B-5 

Monitor ID 

# 

Years 

(3-year 

average) 

Year 

(annual) 

Max 8-hour 

(ppb) 

(3-year 

average) 

Max 8-hour

(ppb) 

(annual) 

W126 

(3-year 

average) 

W126 

(annual) Monitor Site Name State County 

600500021 2010-2012 2010 74 75 17.68 15.56 201 Clinton Road, Jackson CA Amador 600500021 2010-2012 2011 74 70 17.68 14.87 201 Clinton Road, Jackson CA Amador 600500021 2010-2012 2012 74 78 17.68 22.61 201 Clinton Road, Jackson CA Amador 602701011 2008-2010 2008 72 77 17.19 25.85 Death Valley National Monument Near Nevares Springs Access CA Inyo 602701011 2008-2010 2009 72 70 17.19 15.55 Death Valley National Monument Near Nevares Springs Access CA Inyo 602701011 2008-2010 2010 72 69 17.19 10.16 Death Valley National Monument Near Nevares Springs Access CA Inyo 602701011 2009-2011 2009 71 70 16.54 15.55 Death Valley National Monument Near Nevares Springs Access CA Inyo 602701011 2009-2011 2010 71 69 16.54 10.16 Death Valley National Monument Near Nevares Springs Access CA Inyo 602701011 2009-2011 2011 71 75 16.54 23.92 Death Valley National Monument Near Nevares Springs Access CA Inyo 602701011 2010-2012 2010 72 69 18.69 10.16 Death Valley National Monument Near Nevares Springs Access CA Inyo 602701011 2010-2012 2011 72 75 18.69 23.92 Death Valley National Monument Near Nevares Springs Access CA Inyo 602701011 2010-2012 2012 72 73 18.69 22.00 Death Valley National Monument Near Nevares Springs Access CA Inyo 606900031 2005-2007 2005 74 71 15.18 13.11 Pinnacles National Monument, SW of East Entrance Station CA San Benito 606900031 2005-2007 2006 74 78 15.18 17.44 Pinnacles National Monument, SW of East Entrance Station CA San Benito 606900031 2005-2007 2007 74 75 15.18 14.99 Pinnacles National Monument, SW of East Entrance Station CA San Benito 608900071 2008-2010 2009 75 74 15.31 13.66 Anderson - North Street CA Shasta 608900091 2008-2010 2010 75 74 15.31 15.33 Anderson - North Street CA Shasta 608930031 2008-2010 2008 75 83 15.31 18.72 Anderson - North Street CA Shasta 610900051 2009-2011 2009 74 77 20.72 21.80 251 S Barretta, Sonora, CA 95370 CA Tuolumne 610900051 2009-2011 2010 74 72 20.72 20.58 251 S Barretta, Sonora, CA 95370 CA Tuolumne 610900051 2009-2011 2011 74 74 20.72 19.78 251 S Barretta, Sonora, CA 95370 CA Tuolumne 610900051 2010-2012 2010 73 72 20.84 20.58 251 S Barretta, Sonora, CA 95370 CA Tuolumne 610900051 2010-2012 2011 73 74 20.84 19.78 251 S Barretta, Sonora, CA 95370 CA Tuolumne 610900051 2010-2012 2012 73 75 20.84 22.14 251 S Barretta, Sonora, CA 95370 CA Tuolumne 801300111 2000-2002 2000 73 72 15.11 14.06 South Boulder Creek CO Boulder 801300111 2000-2002 2001 73 71 15.11 13.18 South Boulder Creek CO Boulder 801300111 2000-2002 2002 73 78 15.11 18.09 South Boulder Creek CO Boulder 801300111 2003-2005 2003 75 82 16.61 23.91 South Boulder Creek CO Boulder 
5B-6 

Monitor ID 

# 

Years 

(3-year 

average) 

Year 

(annual) 

Max 8-hour 

(ppb) 

(3-year 

average) 

Max 8-hour

(ppb) 

(annual) 

W126 

(3-year 

average) 

W126 

(annual) Monitor Site Name State County 

801300111 2003-2005 2004 75 68 16.61 9.57 South Boulder Creek CO Boulder 801300111 2003-2005 2005 75 76 16.61 16.35 South Boulder Creek CO Boulder 801300111 2004-2006 2004 75 68 17.01 9.57 South Boulder Creek CO Boulder 801300111 2004-2006 2005 75 76 17.01 16.35 South Boulder Creek CO Boulder 801300111 2004-2006 2006 75 82 17.01 25.11 South Boulder Creek CO Boulder 801300111 2008-2010 2008 73 76 16.11 20.77 South Boulder Creek CO Boulder 801300111 2008-2010 2009 73 73 16.11 12.57 South Boulder Creek CO Boulder 801300111 2008-2010 2010 73 72 16.11 14.98 South Boulder Creek CO Boulder 801300111 2009-2011 2009 73 73 16.13 12.57 South Boulder Creek CO Boulder 801300111 2009-2011 2010 73 72 16.13 14.98 South Boulder Creek CO Boulder 801300111 2009-2011 2011 73 76 16.13 20.82 South Boulder Creek CO Boulder 801300111 2010-2012 2010 74 72 19.34 14.98 South Boulder Creek CO Boulder 801300111 2010-2012 2011 74 76 19.34 20.82 South Boulder Creek CO Boulder 801300111 2010-2012 2012 74 76 19.34 22.20 South Boulder Creek CO Boulder 805199911 1998-2000 1998 73 71 20.18 21.13 Gothic CO Gunnison 805199911 1998-2000 1999 73 77 20.18 23.98 Gothic CO Gunnison 805199911 1998-2000 2000 73 73 20.18 15.43 Gothic CO Gunnison 805199911 1999-2001 1999 73 77 18.40 23.98 Gothic CO Gunnison 805199911 1999-2001 2000 73 73 18.40 15.43 Gothic CO Gunnison 805199911 1999-2001 2001 73 70 18.40 15.80 Gothic CO Gunnison 805199911 2000-2002 2000 71 73 18.01 15.43 Gothic CO Gunnison 805199911 2000-2002 2001 71 70 18.01 15.80 Gothic CO Gunnison 805199911 2000-2002 2002 71 71 18.01 22.82 Gothic CO Gunnison 805199911 2001-2003 2001 71 70 18.90 15.80 Gothic CO Gunnison 805199911 2001-2003 2002 71 71 18.90 22.82 Gothic CO Gunnison 805199911 2001-2003 2003 71 73 18.90 18.07 Gothic CO Gunnison 805199911 2002-2004 2002 70 71 17.95 22.82 Gothic CO Gunnison 805199911 2002-2004 2003 70 73 17.95 18.07 Gothic CO Gunnison 805199911 2002-2004 2004 70 67 17.95 12.96 Gothic CO Gunnison 805199911 2003-2005 2003 69 73 15.82 18.07 Gothic CO Gunnison 805199911 2003-2005 2004 69 67 15.82 12.96 Gothic CO Gunnison 805199911 2003-2005 2005 69 69 15.82 16.42 Gothic CO Gunnison 805199911 2004-2006 2004 68 67 15.60 12.96 Gothic CO Gunnison 805199911 2004-2006 2005 68 69 15.60 16.42 Gothic CO Gunnison 805199911 2004-2006 2006 68 70 15.60 17.40 Gothic CO Gunnison 
5B-7 

Monitor ID 

# 

Years 

(3-year 

average) 

Year 

(annual) 

Max 8-hour 

(ppb) 

(3-year 

average) 

Max 8-hour

(ppb) 

(annual) 

W126 

(3-year 

average) 

W126 

(annual) Monitor Site Name State County 

805199911 2005-2007 2005 68 69 16.38 16.42 Gothic CO Gunnison 805199911 2005-2007 2006 68 70 16.38 17.40 Gothic CO Gunnison 805199911 2005-2007 2007 68 65 16.38 15.31 Gothic CO Gunnison 806710041 2005-2007 2005 72 75 18.78 17.93 CO La Plata 806710041 2005-2007 2006 72 74 18.78 20.82 CO La Plata 806710041 2005-2007 2007 72 69 18.78 17.58 CO La Plata 806710041 2006-2008 2006 70 74 18.10 20.82 CO La Plata 806710041 2006-2008 2007 70 69 18.10 17.58 CO La Plata 806710041 2006-2008 2008 70 69 18.10 15.91 CO La Plata 806710041 2008-2010 2008 71 69 15.07 15.91 CO La Plata 806710041 2008-2010 2009 71 71 15.07 10.94 CO La Plata 806710041 2008-2010 2010 71 74 15.07 18.36 CO La Plata 806710041 2009-2011 2009 74 71 16.80 10.94 CO La Plata 806710041 2009-2011 2010 74 74 16.80 18.36 CO La Plata 806710041 2009-2011 2011 74 77 16.80 21.09 CO La Plata 806710041 2010-2012 2010 73 74 19.16 18.36 CO La Plata 806710041 2010-2012 2011 73 77 19.16 21.09 CO La Plata 806710041 2010-2012 2012 73 69 19.16 18.02 CO La Plata 806900071 2008-2010 2010 74 77 18.31 19.12 Fort Collins - West CO Larimer 806900111 2008-2010 2008 74 76 18.31 21.63 Fort Collins - West CO Larimer 806900111 2008-2010 2009 74 73 18.31 14.17 Fort Collins - West CO Larimer 806900071 1999-2001 1999 74 74 15.05 11.16 Rocky Mountain National Park, Long's Peak CO Larimer 806900071 1999-2001 2000 74 78 15.05 25.82 Rocky Mountain National Park, Long's Peak CO Larimer 806900071 1999-2001 2001 74 70 15.05 8.16 Rocky Mountain National Park, Long's Peak CO Larimer 806900071 2004-2006 2004 74 73 15.57 16.23 Rocky Mountain National Park, Long's Peak CO Larimer 806900071 2004-2006 2006 74 76 15.57 18.53 Rocky Mountain National Park, Long's Peak CO Larimer 806999911 2004-2006 2005 74 78 15.57 16.20 Rocky Mountain National Park, Long's Peak CO Larimer 808301011 1998-2000 1998 70 68 16.37 12.90 Mesa Verde National Park, Resource Management Area CO Montezuma 808301011 1998-2000 1999 70 69 16.37 14.17 Mesa Verde National Park, Resource Management Area CO Montezuma 808301011 1998-2000 2000 70 73 16.37 22.04 Mesa Verde National Park, Resource Management Area CO Montezuma 808301011 1999-2001 1999 69 69 15.66 14.17 Mesa Verde National Park, Resource Management Area CO Montezuma 808301011 1999-2001 2000 69 73 15.66 22.04 Mesa Verde National Park, Resource Management Area CO Montezuma 
5B-8 

Monitor ID 

# 

Years 

(3-year 

average) 

Year 

(annual) 

Max 8-hour 

(ppb) 

(3-year 

average) 

Max 8-hour

(ppb) 

(annual) 

W126 

(3-year 

average) 

W126 

(annual) Monitor Site Name State County 

808301011 1999-2001 2001 69 65 15.66 10.77 Mesa Verde National Park, Resource Management Area CO Montezuma 808301011 2000-2002 2000 69 73 17.51 22.04 Mesa Verde National Park, Resource Management Area CO Montezuma 808301011 2000-2002 2001 69 65 17.51 10.77 Mesa Verde National Park, Resource Management Area CO Montezuma 808301011 2000-2002 2002 69 70 17.51 19.72 Mesa Verde National Park, Resource Management Area CO Montezuma 808301011 2001-2003 2001 67 65 16.00 10.77 Mesa Verde National Park, Resource Management Area CO Montezuma 808301011 2001-2003 2002 67 70 16.00 19.72 Mesa Verde National Park, Resource Management Area CO Montezuma 808301011 2001-2003 2003 67 67 16.00 17.50 Mesa Verde National Park, Resource Management Area CO Montezuma 808301011 2002-2004 2002 68 70 16.34 19.72 Mesa Verde National Park, Resource Management Area CO Montezuma 808301011 2002-2004 2003 68 67 16.34 17.50 Mesa Verde National Park, Resource Management Area CO Montezuma 808301011 2002-2004 2004 68 69 16.34 11.79 Mesa Verde National Park, Resource Management Area CO Montezuma 808301011 2003-2005 2003 70 67 16.96 17.50 Mesa Verde National Park, Resource Management Area CO Montezuma 808301011 2003-2005 2004 70 69 16.96 11.79 Mesa Verde National Park, Resource Management Area CO Montezuma 808301011 2003-2005 2005 70 76 16.96 21.59 Mesa Verde National Park, Resource Management Area CO Montezuma 808301011 2004-2006 2004 73 69 19.02 11.79 Mesa Verde National Park, Resource Management Area CO Montezuma 808301011 2004-2006 2005 73 76 19.02 21.59 Mesa Verde National Park, Resource Management Area CO Montezuma 808301011 2004-2006 2006 73 74 19.02 23.68 Mesa Verde National Park, Resource Management Area CO Montezuma 808301011 2005-2007 2005 73 76 21.00 21.59 Mesa Verde National Park, Resource Management Area CO Montezuma 808301011 2005-2007 2006 73 74 21.00 23.68 Mesa Verde National Park, Resource Management Area CO Montezuma 808301011 2005-2007 2007 73 70 21.00 17.73 Mesa Verde National Park, Resource Management Area CO Montezuma 808301011 2006-2008 2006 71 74 18.36 23.68 Mesa Verde National Park, Resource Management Area CO Montezuma 808301011 2006-2008 2007 71 70 18.36 17.73 Mesa Verde National Park, Resource Management Area CO Montezuma 808301011 2006-2008 2008 71 69 18.36 13.67 Mesa Verde National Park, Resource Management Area CO Montezuma 
5B-9 

Monitor ID 

# 

Years 

(3-year 

average) 

Year 

(annual) 

Max 8-hour 

(ppb) 

(3-year 

average) 

Max 8-hour

(ppb) 

(annual) 

W126 

(3-year 

average) 

W126 

(annual) Monitor Site Name State County 

808301011 2007-2009 2007 69 70 15.58 17.73 Mesa Verde National Park, Resource Management Area CO Montezuma 808301011 2007-2009 2008 69 69 15.58 13.67 Mesa Verde National Park, Resource Management Area CO Montezuma 808301011 2007-2009 2009 69 69 15.58 15.35 Mesa Verde National Park, Resource Management Area CO Montezuma 

2106105011 2006-2008 2006 74 71 15.99 12.55 Mammoth Cave National Park, Houchin Meadow KY Edmonson 2106105011 2006-2008 2007 74 82 15.99 22.58 Mammoth Cave National Park, Houchin Meadow KY Edmonson 2106105011 2006-2008 2008 74 70 15.99 12.83 Mammoth Cave National Park, Houchin Meadow KY Edmonson 3501510051 2004-2006 2004 69 65 15.30 8.65 5ZR on BLM Land bordering residential area outside Carlsbad NM Eddy 3501510051 2004-2006 2005 69 67 15.30 10.55 5ZR on BLM Land bordering residential area outside Carlsbad NM Eddy 3501510051 2004-2006 2006 69 76 15.30 26.71 5ZR on BLM Land bordering residential area outside Carlsbad NM Eddy 3501510051 2005-2007 2005 69 67 15.33 10.55 5ZR on BLM Land bordering residential area outside Carlsba NM Eddy 3501510051 2005-2007 2006 69 76 15.33 26.71 5ZR on BLM Land bordering residential area outside Carlsbad NM Eddy 3501510051 2005-2007 2007 69 66 15.33 8.72 5ZR on BLM Land bordering residential area outside Carlsbad NM Eddy 3501510051 2006-2008 2006 69 76 15.07 26.71 5ZR on BLM Land bordering residential area outside Carlsbad NM Eddy 3501510051 2006-2008 2007 69 66 15.07 8.72 5ZR on BLM Land bordering residential area outside Carlsba NM Eddy 3501510051 2006-2008 2008 69 67 15.07 9.78 5ZR on BLM Land bordering residential area outside Carlsbad NM Eddy 3504310011 2000-2002 2001 72 69 17.19 12.17 NM Sandoval 3504310011 2000-2002 2002 72 74 17.19 19.62 NM Sandoval 3504310011 2001-2003 2001 71 69 17.37 12.17 NM Sandoval 3504310011 2001-2003 2002 71 74 17.37 19.62 NM Sandoval 3504310031 2000-2002 2000 72 75 17.19 23.51 NM Sandoval 3504310031 2001-2003 2003 71 76 17.37 25.38 NM Sandoval 3504310011 1999-2001 2001 72 69 17.87 12.17 2ZR Site moved from Rio Rancho City Hall to senior center NM Sandoval 3504310011 2002-2004 2002 74 74 20.86 19.62 2ZR Site moved from Rio Rancho City Hall to senior center NM Sandoval 3504310011 2002-2004 2004 74 71 20.86 17.73 2ZR Site moved from Rio Rancho City Hall to senior center NM Sandoval 
5B-10 

Monitor ID 

# 

Years 

(3-year 

average) 

Year 

(annual) 

Max 8-hour 

(ppb) 

(3-year 

average) 

Max 8-hour

(ppb) 

(annual) 

W126 

(3-year 

average) 

W126 

(annual) Monitor Site Name State County 

3504310011 2003-2005 2004 74 71 20.08 17.73 2ZR Site moved from Rio Rancho City Hall to senior center NM Sandoval 3504310011 2004-2006 2004 73 71 17.75 17.73 2ZR Site moved from Rio Rancho City Hall to senior center NM Sandoval 3504310031 1999-2001 1999 72 76 17.87 18.14 2ZR Site moved from Rio Rancho City Hall to senior center NM Sandoval 3504310031 1999-2001 2000 72 75 17.87 23.51 2ZR Site moved from Rio Rancho City Hall to senior center NM Sandoval 3504310031 2002-2004 2003 74 76 20.86 25.38 2ZR Site moved from Rio Rancho City Hall to senior center NM Sandoval 3504310031 2003-2005 2003 74 76 20.08 25.38 2ZR Site moved from Rio Rancho City Hall to senior center NM Sandoval 3504310031 2003-2005 2005 74 75 20.08 17.03 2ZR Site moved from Rio Rancho City Hall to senior center NM Sandoval 3504310031 2004-2006 2005 73 75 17.75 17.03 2ZR Site moved from Rio Rancho City Hall to senior center NM Sandoval 3504310031 2005-2007 2005 73 75 17.50 17.03 2ZR Site moved from Rio Rancho City Hall to senior center NM Sandoval 3504310031 2005-2007 2007 73 71 17.50 17.05 2ZR Site moved from Rio Rancho City Hall to senior center NM Sandoval 3504310031 2006-2008 2007 70 71 15.87 17.05 2ZR Site moved from Rio Rancho City Hall to senior center NM Sandoval 3504310031 2006-2008 2008 70 65 15.87 12.15 2ZR Site moved from Rio Rancho City Hall to senior center NM Sandoval 3504390041 2004-2006 2006 73 72 17.75 19.26 2ZR Site moved from Rio Rancho City Hall to senior center NM Sandoval 3504390041 2005-2007 2006 73 72 17.50 19.26 2ZR Site moved from Rio Rancho City Hall to senior center NM Sandoval 3504390041 2006-2008 2006 70 72 15.87 19.26 2ZR Site moved from Rio Rancho City Hall to senior center NM Sandoval 4603301323 2005-2007 2005 70 70 15.49 13.56 Wind Cave National Park, Visitor Center SD Custer 4603301323 2005-2007 2006 70 73 15.49 20.63 Wind Cave National Park, Visitor Center SD Custer 4603301323 2005-2007 2007 70 69 15.49 12.29 Wind Cave National Park, Visitor Center SD Custer 4903701011 1998-2000 1998 73 71 19.80 19.78 Canyonlands National Park, Island in the Sky UT San Juan 4903701011 1998-2000 1999 73 73 19.80 20.25 Canyonlands National Park, Island in the Sky UT San Juan 4903701011 1998-2000 2000 73 76 19.80 19.36 Canyonlands National Park, Island in the Sky UT San Juan 4903701011 2001-2003 2001 70 66 18.94 9.91 Canyonlands National Park, Island in the Sky UT San Juan 4903701011 2001-2003 2002 70 72 18.94 22.12 Canyonlands National Park, Island in the Sky UT San Juan 4903701011 2001-2003 2003 70 74 18.94 24.80 Canyonlands National Park, Island in the Sky UT San Juan 4903701011 2002-2004 2002 72 72 20.50 22.12 Canyonlands National Park, Island in the Sky UT San Juan 4903701011 2002-2004 2003 72 74 20.50 24.80 Canyonlands National Park, Island in the Sky UT San Juan 
5B-11 

Monitor ID 

# 

Years 

(3-year 

average) 

Year 

(annual) 

Max 8-hour 

(ppb) 

(3-year 

average) 

Max 8-hour

(ppb) 

(annual) 

W126 

(3-year 

average) 

W126 

(annual) Monitor Site Name State County 

4903701011 2002-2004 2004 72 72 20.50 14.57 Canyonlands National Park, Island in the Sky UT San Juan 4903701011 2003-2005 2003 71 74 18.59 24.80 Canyonlands National Park, Island in the Sky UT San Juan 4903701011 2003-2005 2004 71 72 18.59 14.57 Canyonlands National Park, Island in the Sky UT San Juan 4903701011 2003-2005 2005 71 69 18.59 16.40 Canyonlands National Park, Island in the Sky UT San Juan 4903701011 2004-2006 2004 70 72 16.59 14.57 Canyonlands National Park, Island in the Sky UT San Juan 4903701011 2004-2006 2005 70 69 16.59 16.40 Canyonlands National Park, Island in the Sky UT San Juan 4903701011 2004-2006 2006 70 70 16.59 18.80 Canyonlands National Park, Island in the Sky UT San Juan 4903701011 2005-2007 2005 70 69 17.66 16.40 Canyonlands National Park, Island in the Sky UT San Juan 4903701011 2005-2007 2006 70 70 17.66 18.80 Canyonlands National Park, Island in the Sky UT San Juan 4903701011 2005-2007 2007 70 72 17.66 17.78 Canyonlands National Park, Island in the Sky UT San Juan 4903701011 2006-2008 2006 71 70 18.10 18.80 Canyonlands National Park, Island in the Sky UT San Juan 4903701011 2006-2008 2007 71 72 18.10 17.78 Canyonlands National Park, Island in the Sky UT San Juan 4903701011 2006-2008 2008 71 71 18.10 17.71 Canyonlands National Park, Island in the Sky UT San Juan 4903701011 2007-2009 2007 70 72 16.07 17.78 Canyonlands National Park, Island in the Sky UT San Juan 4903701011 2007-2009 2008 70 71 16.07 17.71 Canyonlands National Park, Island in the Sky UT San Juan 4903701011 2007-2009 2009 70 68 16.07 12.72 Canyonlands National Park, Island in the Sky UT San Juan 4903701011 2010-2012 2010 69 68 15.01 13.87 Canyonlands National Park, Island in the Sky UT San Juan 4903701011 2010-2012 2011 69 69 15.01 14.23 Canyonlands National Park, Island in the Sky UT San Juan 4903701011 2010-2012 2012 69 72 15.01 16.93 Canyonlands National Park, Island in the Sky UT San Juan 4905301301 2006-2008 2006 71 72 21.12 24.29 Zion National Park, Dalton's Wash UT Washington 4905301301 2006-2008 2007 71 71 21.12 19.37 Zion National Park, Dalton's Wash UT Washington 4905301301 2006-2008 2008 71 72 21.12 19.69 Zion National Park, Dalton's Wash UT Washington 4905301301 2007-2009 2007 70 71 18.01 19.37 Zion National Park, Dalton's Wash UT Washington 4905301301 2007-2009 2008 70 72 18.01 19.69 Zion National Park, Dalton's Wash UT Washington 4905301301 2007-2009 2009 70 68 18.01 14.96 Zion National Park, Dalton's Wash UT Washington 4905301301 2008-2010 2008 70 72 18.46 19.69 Zion National Park, Dalton's Wash UT Washington 4905301301 2008-2010 2009 70 68 18.46 14.96 Zion National Park, Dalton's Wash UT Washington 4905301301 2008-2010 2010 70 72 18.46 20.73 Zion National Park, Dalton's Wash UT Washington 4905301301 2009-2011 2009 70 68 17.85 14.96 Zion National Park, Dalton's Wash UT Washington 4905301301 2009-2011 2010 70 72 17.85 20.73 Zion National Park, Dalton's Wash UT Washington 4905301301 2009-2011 2011 70 72 17.85 17.86 Zion National Park, Dalton's Wash UT Washington 4905301301 2010-2012 2010 73 72 20.34 20.73 Zion National Park, Dalton's Wash UT Washington 4905301301 2010-2012 2011 73 72 20.34 17.86 Zion National Park, Dalton's Wash UT Washington 4905301301 2010-2012 2012 73 75 20.34 22.42 Zion National Park, Dalton's Wash UT Washington 5603599911 1998-2000 1998 72 71 17.25 16.02 Pinedale WY Sublette 
5B-12 

Monitor ID 

# 

Years 

(3-year 

average) 

Year 

(annual) 

Max 8-hour 

(ppb) 

(3-year 

average) 

Max 8-hour

(ppb) 

(annual) 

W126 

(3-year 

average) 

W126 

(annual) Monitor Site Name State County 

5603599911 1998-2000 1999 72 72 17.25 16.88 Pinedale WY Sublette 5603599911 1998-2000 2000 72 73 17.25 18.86 Pinedale WY Sublette 5603599911 1999-2001 1999 71 72 16.68 16.88 Pinedale WY Sublette 5603599911 1999-2001 2000 71 73 16.68 18.86 Pinedale WY Sublette 5603599911 1999-2001 2001 71 69 16.68 14.31 Pinedale WY Sublette 5603599911 2000-2002 2000 71 73 17.46 18.86 Pinedale WY Sublette 5603599911 2000-2002 2001 71 69 17.46 14.31 Pinedale WY Sublette 5603599911 2000-2002 2002 71 72 17.46 19.21 Pinedale WY Sublette 5603599911 2001-2003 2001 70 69 16.63 14.31 Pinedale WY Sublette 5603599911 2001-2003 2002 70 72 16.63 19.21 Pinedale WY Sublette 5603599911 2001-2003 2003 70 70 16.63 16.36 Pinedale WY Sublette 5603599911 2002-2004 2002 69 72 15.16 19.21 Pinedale WY Sublette 5603599911 2002-2004 2003 69 70 15.16 16.36 Pinedale WY Sublette 5603599911 2002-2004 2004 69 65 15.16 9.93 Pinedale WY Sublette 
5C-1 

APPENDIX 5C: EXPANDED EVALUATION OF RELATIVE BIOMASS 

AND YIELD LOSS 

This appendix expands to range W126 index values evaluated for relative biomass and 

yield loss. Specifically, Tables 5C-1 and 5C-2 below provide estimates of the relative loss for 

trees and crops respectively at various W126 index values using the composite E-R functions for 

each species for each integer W126 index value between 7 ppm-hrs and 30 ppm-hrs. The median 

of the composite functions is calculated for all 11 tree species excluding cottonwood. These 

tables also provide estimates of the number of species for trees and crops respectively that would 

be below various benchmarks (e.g., 2% biomass loss for trees) at various W126 index values. 

Table 5C-3 provides an expansion of Table 6-1 to reflect each integer W126 index value 

between 7 ppm-hrs and 23 ppm-hrs.
5C-2 

Table 5C-1 Relative Biomass Loss for Eleven Individual Tree Seedlings and Median at Various W126 Index Valu

W126 Douglas Fir Loblolly Virginia Pine Red maple Sugar maple Red Alder Ponderosa Pine Aspen Tulip Poplar Eastern White Pine Black Cherry Median (11 species) Number of Species ≤ 2% 

o

30 0.1% 0.8% 1.7% 3.8% 28.1% 10.4% 12.8% 18.6% 27.7% 25.2% 53.8% 12.8% 3 29 0.0% 0.7% 1.7% 3.6% 23.7% 10.0% 12.3% 17.9% 26.1% 24.0% 52.6% 12.3% 3 28 0.0% 0.7% 1.6% 3.5% 19.9% 9.6% 11.8% 17.2% 24.5% 22.8% 51.4% 11.8% 3 27 0.0% 0.7% 1.6% 3.3% 16.4% 9.2% 11.4% 16.5% 23.0% 21.6% 50.1% 11.4% 3 26 0.0% 0.7% 1.5% 3.1% 13.4% 8.8% 10.9% 15.8% 21.4% 20.5% 48.8% 10.9% 3 25 0.0% 0.6% 1.4% 3.0% 10.9% 8.4% 10.4% 15.2% 19.9% 19.3% 47.5% 10.4% 3 24 0.0% 0.6% 1.4% 2.8% 8.7% 8.0% 10.0% 14.5% 18.4% 18.2% 46.2% 8.7% 3 23 0.0% 0.6% 1.3% 2.7% 6.9% 7.6% 9.5% 13.8% 17.0% 17.1% 44.8% 7.6% 3 22 0.0% 0.6% 1.3% 2.5% 5.3% 7.2% 9.0% 13.1% 15.6% 15.9% 43.3% 7.2% 3 21 0.0% 0.5% 1.2% 2.4% 4.1% 6.8% 8.6% 12.4% 14.3% 14.9% 41.9% 6.8% 3 20 0.0% 0.5% 1.2% 2.2% 3.1% 6.4% 8.1% 11.8% 13.0% 13.8% 40.3% 6.4% 3 19 0.0% 0.5% 1.1% 2.1% 2.3% 6.0% 7.6% 11.1% 11.8% 12.7% 38.8% 6.0% 3 18 0.0% 0.5% 1.0% 1.9% 1.7% 5.7% 7.2% 10.4% 10.6% 11.7% 37.2% 5.7% 5 17 0.0% 0.4% 1.0% 1.8% 1.2% 5.3% 6.7% 9.8% 9.4% 10.7% 35.6% 5.3% 5 16 0.0% 0.4% 0.9% 1.6% 0.9% 4.9% 6.3% 9.1% 8.4% 9.7% 33.9% 4.9% 5 15 0.0% 0.4% 0.9% 1.5% 0.6% 4.5% 5.8% 8.4% 7.4% 8.8% 32.2% 4.5% 5 14 0.0% 0.4% 0.8% 1.4% 0.4% 4.2% 5.4% 7.8% 6.4% 7.9% 30.4% 4.2% 5 13 0.0% 0.3% 0.8% 1.2% 0.3% 3.8% 4.9% 7.1% 5.5% 7.0% 28.6% 3.8% 5 12 0.0% 0.3% 0.7% 1.1% 0.2% 3.5% 4.5% 6.5% 4.7% 6.2% 26.7% 3.5% 5 11 0.0% 0.3% 0.6% 1.0% 0.1% 3.1% 4.1% 5.9% 3.9% 5.4% 24.8% 3.1% 5 10 0.0% 0.3% 0.6% 0.9% 0.1% 2.8% 3.6% 5.2% 3.2% 4.6% 22.9% 2.8% 5 

9 0.0% 0.2% 0.5% 0.7% 0.0% 2.4% 3.2% 4.6% 2.6% 3.9% 20.9% 2.4% 5 8 0.0% 0.2% 0.5% 0.6% 0.0% 2.1% 2.8% 4.0% 2.0% 3.2% 18.8% 2.0% 5 7 0.0% 0.2% 0.4% 0.5% 0.0% 1.8% 2.4% 3.4% 1.5% 2.6% 16.7% 1.5% 7 
5C-3 

Table 5C-2 Relative Yield Loss for Ten Individual Crop Species and Median at Various W126 Index Values

W126 Barley Lettuce Field Corn Grain Sorghum Peanut Cotton Soybean Winter Wheat Potato Kidney Bean Median (10 species) Number of Species ≤ 5% Number of Species ≤ 10% Numb

of Spec

≤ 20% 

30 0.1% 5.1% 2.9% 2.3% 10.4% 16.3% 15.7% 22.5% 20.2% 36.1% 13.0% 3 4 7 29 0.0% 4.4% 2.7% 2.1% 9.7% 15.6% 15.0% 21.0% 19.4% 34.0% 12.4% 4 5 8 28 0.0% 3.7% 2.4% 2.0% 9.1% 14.9% 14.4% 19.5% 18.7% 31.9% 11.8% 4 5 9 27 0.0% 3.1% 2.2% 1.9% 8.6% 14.1% 13.7% 18.0% 18.0% 29.8% 11.2% 4 5 9 26 0.0% 2.6% 1.9% 1.7% 8.0% 13.4% 13.1% 16.6% 17.2% 27.8% 10.6% 4 5 9 25 0.0% 2.1% 1.7% 1.6% 7.4% 12.7% 12.5% 15.3% 16.5% 25.8% 10.0% 4 5 9 24 0.0% 1.7% 1.5% 1.5% 6.9% 12.0% 11.8% 14.0% 15.7% 23.9% 9.4% 4 5 9 23 0.0% 1.4% 1.3% 1.4% 6.4% 11.3% 11.2% 12.7% 15.0% 22.0% 8.8% 4 5 9 22 0.0% 1.1% 1.2% 1.3% 5.9% 10.6% 10.6% 11.5% 14.2% 20.1% 8.2% 4 5 9 21 0.0% 0.9% 1.0% 1.1% 5.4% 10.0% 10.0% 10.4% 13.5% 18.4% 7.7% 4 7 10 20 0.0% 0.7% 0.9% 1.0% 5.0% 9.3% 9.4% 9.3% 12.7% 16.6% 7.1% 5 8 10 19 0.0% 0.6% 0.8% 0.9% 4.5% 8.7% 8.8% 8.3% 12.0% 15.0% 6.4% 5 8 10 18 0.0% 0.4% 0.7% 0.8% 4.1% 8.0% 8.2% 7.3% 11.3% 13.4% 5.7% 5 8 10 17 0.0% 0.3% 0.6% 0.8% 3.7% 7.4% 7.6% 6.4% 10.5% 11.9% 5.1% 5 8 10 16 0.0% 0.2% 0.5% 0.7% 3.3% 6.8% 7.0% 5.6% 9.8% 10.5% 4.4% 5 9 10 15 0.0% 0.2% 0.4% 0.6% 2.9% 6.2% 6.4% 4.8% 9.1% 9.2% 3.9% 6 10 10 14 0.0% 0.1% 0.3% 0.5% 2.6% 5.6% 5.9% 4.1% 8.4% 7.9% 3.3% 6 10 10 13 0.0% 0.1% 0.2% 0.5% 2.2% 5.0% 5.3% 3.5% 7.7% 6.8% 2.8% 6 10 10 12 0.0% 0.1% 0.2% 0.4% 1.9% 4.5% 4.8% 2.9% 7.0% 5.7% 2.4% 8 10 10 11 0.0% 0.0% 0.2% 0.3% 1.6% 3.9% 4.3% 2.3% 6.3% 4.7% 2.0% 9 10 10 10 0.0% 0.0% 0.1% 0.3% 1.4% 3.4% 3.8% 1.9% 5.6% 3.8% 1.6% 9 10 10 

9 0.0% 0.0% 0.1% 0.2% 1.1% 2.9% 3.3% 1.5% 4.9% 3.0% 1.3% 10 10 10 8 0.0% 0.0% 0.1% 0.2% 0.9% 2.5% 2.8% 1.1% 4.3% 2.4% 1.0% 10 10 10 7 0.0% 0.0% 0.0% 0.1% 0.7% 2.0% 2.3% 0.8% 3.6% 1.8% 0.8% 10 10 10 
5C-4 

Table 5C-3 Tree Seedling Biomass Loss and Crop Yield Loss estimated for O3 exposure 

over a Season. 

W126 index 

value 

for exposure 

period 

Tree seedling biomass lossA Crop yield lossC

Median Value Individual Species Median Value Individual Species 

23 ppm-hrs Median species 

w. 7.6% loss B

< 2% loss: 3/11 species 

< 5% loss: 4/11 species 

<10% loss: 8/11 species 

<15% loss: 10/11 species 

>40% loss: 1/11 species 

Median species w. 

8.8 % loss D

< 5% loss: 4/10 species 

>5,<10% loss: 1/10 species 

>10,<20% loss: 4/10 species 

>20: 1/10 species 

22 ppm-hrs Median species 

w. 7.2% loss B

< 2% loss: 3/11 species 

< 5% loss: 4/11 species 

<10% loss: 7/11 species 

<15% loss: 10/11 species 

>40% loss: 1/11 species 

Median species w. 

8.2 % loss D

< 5% loss: 4/10 species 

>5,<10% loss: 1/10 species 

>10,<20% loss: 4/10 species 

>20: 1/10 species 

21 ppm-hrs Median species 

w. 6.8% loss B

< 2% loss: 3/11 species 

< 5% loss: 4/11 species 

<10% loss: 7/11 species 

<15% loss: 10/11 species 

>40% loss: 1/11 species 

Median species w. 

7.7 % loss D

< 5% loss: 4/10 species 

>5,<10% loss: 3/10 species 

>10,<20% loss: 3/10 species 

20 ppm-hrs Median species 

w. 6.4% loss B

< 2% loss: 3/11 species 

< 5% loss: 5/11 species 

<10% loss: 7/11 species 

<15% loss: 10/11 species 

>40% loss: 1/11 species

Median species w. 

7.1 % loss D

< 5% loss: 5/10 species 

>5,<10% loss: 3/10 species 

>10,<20% loss: 2/10 species 

19 ppm-hrs Median species 

w. 6.0% loss B

< 2% loss: 3/11 species 

<5% loss: 5/11 species 

<10% loss: 7/11 species 

<15% loss: 10/11 species 

>30% loss: 1/11 species 

Median species w. 

6.4 % loss D

< 5% loss: 5/10 species 

>5, <10% loss: 3/10 species 

>10,<20% loss: 2/10 species 

18 ppm-hrs Median species 

w. 5.7% loss B

< 2% loss: 5/11 species 

< 5% loss: 5/11 species 

<10% loss: 7/11 species 

<15% loss: 10/11 species 

>30% loss: 1/11 species

Median species w. 

5.7 % loss D

< 5% loss: 5/10 species 

>5,<10% loss: 3/10 species 

>10,<20% loss: 2/10 species 

17 ppm-hrs Median species 

w. 5.3% loss B

< 2% loss: 5/11 species 

<5% loss: 5/11 species 

<10% loss: 9/11 species 

<15% loss: 10/11 species 

>30% loss: 1/11 species 

Median species w. 

5.1 % loss D

< 5% loss: 5/10 species 

>5, <10% loss: 3/10 species 

>10,<20% loss: 2/10 species 

16 ppm-hrs Median species 

w. 4.9% loss B

< 2% loss: 5/11 species 

< 5% loss: 6/11 species 

<10% loss: 10/11 species 

>30% loss: 1/11 species

Median species w. 

4.4 % loss D

< 5% loss: 5/10 species 

>5,<10% loss: 4/10 species 

>10,<20% loss: 1/10 species 

15 ppm-hrs Median species 

w. 4.5% loss B

< 2% loss: 5/11 species 

<5% loss: 6/11 species 

<10% loss: 10/11 species 

>30% loss: 1/11 species 

Median species w. 

<5% loss D

< 5% loss: 6/10 species 

>5, <10% loss: 4/10 species 
5C-5 

14 ppm-hrs Median species 

w. 4.2% loss B

< 2% loss: 5/11 species 

< 5% loss: 6/11 species 

<10% loss: 10/11 species 

>30% loss: 1/11 species

Median species w. 

<5% loss D

< 5% loss: 6/10 species 

>5,<10% loss: 4/10 species 

13 ppm-hrs Median species 

w. 3.8% loss B

< 2% loss: 5/11 species 

<5% loss: 7/11 species 

<10% loss: 10/11 species 

>20% loss: 1/11 species 

Median species 

w.<5% loss D

< 5% loss: 6/10 species 

>5, <10% loss: 4/10 species 

12 ppm-hrs Median species 

w. 3.5% loss B

< 2% loss: 5/11 species 

< 5% loss: 8/11 species 

<10% loss: 10/11 species 

>20% loss: 1/11 species

Median species w. 

<5% loss D

< 5% loss: 8/10 species 

>5,<10% loss: 2/10 species 

11 ppm-hrs Median species 

w. 3.1% loss B

< 2% loss: 5/11 species 

<5% loss: 8/11 species 

<10% loss: 10/11 species 

>20% loss: 1/11 species 

Median species w. 

<5% loss D

< 5% loss: 9/10 species 

>5, <10% loss: 1/10 species 

10 ppm-hrs Median species 

w. 2.8% loss B

< 2% loss: 5/11 species 

< 5% loss: 9/11 species 

<10% loss: 10/11 species 

>20% loss: 1/11 species 

Median species w. 

<5% loss D

< 5% loss: 9/10 species 

>5,<10% loss: 1/10 species 

9 ppm-hrs Median species 

w. 2.4% loss B

< 2% loss: 5/11 species 

< 5% loss: 10/11 species 

>20% loss: 1/11 species 

Median species w. 

<5% loss D

< 5% loss: all species 

8 ppm-hrs Median species 

w. 2.0% loss B

< 2% loss: 5/11 species 

< 5% loss: 10/11 species 

>15% loss: 1/11 species 

Median species w. 

<5% loss D < 5% loss: all species 

7 ppm-hrs Median species 

w. <2% loss B

< 2% loss: 7/11 species 

<5% loss: 10/11 species 

>15% loss: 1/11 species 

Median species w. 

<5% loss D

< 5% loss: all species 

A Estimates here are based on the 11 E-R functions for tree seedlings described in WREA, Appendix 6F and 

discussed in section 5.2.1, with the exclusion of cottonwood. See CASAC comments (Frey, 2014). 

B This median value is the median of the composite E-R functions for 11 tree species in the WREA, Appendix 6F 

(also discussed in section 5.2.1). 

C Estimates here are based on the 10 E-R functions for crops described in Appendix 6F and discussed in section 

5.3.1. 

D This median value is the median of the composite E-R functions for 10 crops from WREA, Appendix 6F (also 

discussed in section 5.3.1). 
United States 

Environmental Protection Agency 

Office of Air Quality Planning and Standards 

Health and Environmental Impacts Division 

Research Triangle Park, NC 

Publication N