Policy Assessment for the Review of the 

Ozone National Ambient Air Quality 

Standards 

Second External Review Draft 
 

DISCLAIMER

This draft 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 preliminary conclusions of the staff of the OAQPS and does 

not necessarily reflect the views of the Agency. This draft document is being circulated to 

facilitate discussion with the Clean Air Scientific Advisory Committee to inform the EPA’s 

consideration of the ozone National Ambient Air Quality Standards. 

This information is distributed for the purposes of pre-dissemination peer review under 

applicable information quality guidelines. It has not been formally disseminated by EPA. It 

does not represent and should not be construed to represent any Agency determination or policy. 

Mention of trade names or commercial products is not intended to constitute endorsement or 

recommendation for use. 
EPA–452/P–14–002 

January 2014 

Policy Assessment for the Review of the Ozone National Ambient Air 

Quality Standards 

Second External Review Draft

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

1 TABLE OF CONTENTS 

2 LISTS OF FIGURES..............................................................................vi 

3 LIST OF TABLES..................................................................................x

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

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

6 1.1 PURPOSE ....................................................................................................... 1-1

7 1.2 BACKGROUND ............................................................................................. 1-3

8 1.2.1 Legislative Requirements ......................................................................... 1-3

9 1.2.2 History of O3 NAAQS Reviews ............................................................... 1-5

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

11 1.3 GENERAL APPROACH FOR REVIEW OF THE STANDARDS ............. 1-12

12 1.3.1 Approach for the Primary Standard ....................................................... 1-13

13 1.3.1.1 Approach Used in the Last Review ............................................... 1-14

14 1.3.1.2 Approach for the Current Review ................................................. 1-17

15 1.3.1.2.1 Consideration of the Scientific Evidence ..................................... 1-20

16 1.3.1.2.2 Consideration of Exposure and Risk Estimates ........................... 1-24

17 1.3.1.2.3 Considerations Regarding Ambient O3 Concentration Estimates 

18 Attributable to Background Sources ........................................... 1-25

19 1.3.2 Approach for the Secondary Standard ................................................... 1-26

20 1.3.2.1 Approach Used in the Last Review ............................................... 1-27

21 1.3.2.2 Approach for the Current Review ................................................. 1-32

22 1.3.2.2.1 Consideration of the Scientific Evidence ..................................... 1-35

23 1.3.2.2.2 Consideration of Exposure and Risk Estimates and Air Quality 

24 Analyses ...................................................................................... 1-38

25 1.3.2.2.3 Considerations Regarding Ambient O3 Concentration Estimates 

26 Attributable to Background Sources ........................................... 1-40

27 1.3.3 Organization of this Document .............................................................. 1-40

28 1.4 REFERENCES .............................................................................................. 1-41

29 2 O3 MONITORING AND AIR QUALITY ............................................................... 2-1

30 2.1 O3 MONITORING........................................................................................... 2-1

31 2.1.1 O3 Monitoring Network ............................................................................. 2-1
ii

1 2.1.2 Recent O3 Monitoring Data and Trends..................................................... 2-2

2 2.2 EMISSIONS AND ATMOSPHERIC CHEMISTRY ..................................... 2-7

3 2.3 AIR QUALITY CONCENTRATIONS ........................................................... 2-9

4 2.4 BACKGROUND O3 ...................................................................................... 2-11

5 2.4.1 Seasonal Mean Background O3 in the U.S. ............................................. 2-13

6 2.4.2 Seasonal Mean Background O3 in the U.S. as a Proportion of Total O3 . 2-15

7 2.4.3 Daily Distributions of Background O3 within the Seasonal Mean .......... 2-16

8 2.4.4 Proportion of Background O3 in 12 Urban Case Study Areas ................. 2-20

9 2.4.5 Influence of Background O3 on W126 levels .......................................... 2-22

10 2.4.6 Estimated Magnitude of Individual Components of Background O3 ...... 2-24

11 2.4.7 Summary .................................................................................................. 2-26

12 2.5 REFERENCES .............................................................................................. 2-27

13 3 ADEQUACY OF THE CURRENT PRIMARY STANDARD.................3-1

14 3.1 EVIDENCE-BASED CONSIDERATIONS ................................................... 3-1

15 3.1.1 Modes of Action ........................................................................................ 3-2

16 3.1.2 Nature of Effects ........................................................................................ 3-5

17 3.1.2.1 Respiratory Effects – Short-term Exposures..................................... 3-6

18 3.1.2.3 Total Mortality – Short-term Exposures ......................................... 3-46

19 3.1.2.4 Cardiovascular effects – Short-term Exposure ............................... 3-51

20 3.1.3 Adversity of Effects ................................................................................. 3-54

21 3.1.4 Ozone Concentrations Associated With Health Effects .......................... 3-58

22 3.1.4.1 Concentrations in Controlled Human Exposure Studies and in 

23 Epidemiologic Panel Studies ....................................................................... 3-58

24 3.1.4.2 Concentrations in Epidemiologic Studies – Short-term Metrics .... 3-61

25 3.1.4.3 Concentrations in Epidemiologic Studies – “Long-term” Metrics . 3-75

26 3.1.5 Public Health Implications ....................................................................... 3-78

27 3.1.5.1 At-Risk Populations ........................................................................ 3-78

28 3.1.5.2 Size of At-Risk Populations and Lifestages in the United States ... 3-81

29 3.1.5.3 Averting Behavior ........................................................................... 3-83

30 3.2 AIR QUALITY-, EXPOSURE-, AND RISK-BASED CONSIDERATIONS ..... 

31 ..........................................................................................3-84

32 3.2.1 Consideration of the Adjusted Air Quality Used in Exposure and Risk 

33 Assessments ............................................................................................. 3-84

34 3.2.2 Exposure-Based Considerations .............................................................. 3-87
iii

1 3.2.3 Risk-Based Considerations ...................................................................... 3-97

2 3.2.3.1 Risk of Lung Function Decrements ................................................ 3-97

3 3.2.3.2 Estimated Health Risks Associated with Short- or Long-Term O3

4 Exposures, Based on Epidemiologic Studies ............................... 3-106

5 3.3 CASAC ADVICE ........................................................................................ 3-115

6 3.4 PRELIMINARY STAFF CONCLUSIONS ON ADEQUACY OF PRIMARY 

7 STANDARD ............................................................................................... 3-117

8 3.5 REFERENCES ............................................................................................ 3-124

9 4 CONSIDERATION OF ALTERNATIVE PRIMARY STANDARDS ............... 4-1

10 4.1 INDICATOR.................................................................................................... 4-1

11 4.2 AVERAGING TIME ....................................................................................... 4-2

12 4.3 FORM .............................................................................................................. 4-5

13 4.4 LEVEL ............................................................................................................. 4-8

14 4.4.1 Evidence-based Considerations................................................................ 4-8

15 4.4.2 Air Quality-, Exposure-, and Risk-Based Considerations ..................... 4-19

16 4.4.2.1 Exposure-Based Considerations .................................................... 4-19

17 4.4.2.2 Risk-Based Considerations: Lung Function .................................. 4-27

18 4.4.2.3 Risk-Based Considerations: Epidemiology-Based Mortality and 

19 Morbidity ...................................................................................... 4-34

20 4.5 CASAC ADVICE .......................................................................................... 4-42

21 4.6 PRELIMINARY STAFF CONCLUSIONS ON ALTERNATIVE PRIMARY 

22 STANDARDS FOR CONSIDERATION ..................................................... 4-43

23 4.7 KEY UNCERTAINTIES AND AREAS FOR FUTURE RESEARCH AND 

24 DATA COLLECTION .................................................................................. 4-57

25 4.8 SUMMARY OF PRELIMINARY STAFF CONCLUSIONS ON PRIMARY 

26 STANDARD ................................................................................................. 4-60

27 4.9 REFERENCES........................................................................4-64

28 5 ADEQUACY OF THE CURRENT SECONDARY STANDARD..................5-1

29 5.1 NATURE OF EFFECTS AND BIOLOGICALLY-RELEVANT 

30 EXPOSURE METRIC .................................................................................... 5-1

31 5.2 FOREST TREE GROWTH, PRODUCTIVITY AND CARBON 

32 STORAGE ....................................................................................................... 5-7

33 5.2.1 Evidence-based Considerations................................................................ 5-7

34 5.2.2 Exposure/Risk-based Considerations ..................................................... 5-20

35 5.3 CROP YIELD LOSS ..................................................................................... 5-29
iv

1 5.3.1 Evidence-based Considerations.............................................................. 5-29

2 5.3.2 Exposure/Risk-based Considerations ..................................................... 5-32

3 5.4 VISIBLE FOLIAR INJURY ......................................................................... 5-35

4 5.4.1 Evidence-based Considerations.............................................................. 5-37

5 5.4.2 Exposure-/Risk-based Considerations.................................................... 5-44

6 5.5 OTHER WELFARE EFFECTS .................................................................... 5-54

7 5.5.1 Forest Susceptibility to Insect Infestation .............................................. 5-54

8 5.5.2 Fire Regulation ....................................................................................... 5-55

9 5.5.3 Ozone Effects on Climate ...................................................................... 5-56

10 5.5.4 Additional Effects .................................................................................. 5-57

11 5.6 CASAC ADVICE .......................................................................................... 5-57

12 5.7 PRELIMINARY STAFF CONCLUSIONS ON ADEQUACY OF 

13 SECONDARY STANDARD ........................................................................ 5-59

14 5.8 REFERENCES .............................................................................................. 5-65

15 6 CONSIDERATION OF ALTERNATIVE SECONDARY STANDARDS ......... 6-1

16 6.1 INDICATOR.................................................................................................... 6-1

17 6.2 FORM AND AVERAGING TIME ................................................................. 6-2

18 6.3 LEVEL ........................................................................................................... 6-16

19 6.4 CASAC ADVICE .......................................................................................... 6-34

20 6.5 PRELIMINARY STAFF CONCLUSIONS ON ALTERNATIVE 

21 STANDARD .................................................................................................. 6-36

22 6.6 SUMMARY OF PRELIMINARY CONCLUSIONS ON THE 

23 SECONDARY STANDARD ........................................................................ 6-45

24 6.7 KEY UNCERTAINTIES AND AREAS FOR FUTURE RESEARCH 

25 AND DATA COLLECTION ......................................................................... 6-47

26 6.8 REFERENCES........................................................................6-49

27 APPENDICES 

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

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

30 W126 Metric.............................................................................2B-1 

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

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

33 Appendix 3A. Modes of Action Summary ......................................................................... 3A-1 
v

1 Appendix 3B. Recent Studies of Respiratory-related Emergency Department Visits and 

2 Hospital Admissions....................................................................3B-1 

3 Appendix 3C. At-risk Populations. ......................................................................................3C-1

4 Appendix 3D. Air Quality Data for Locations of Key Epidemiological Studies .............. 3D-1 

5 Appendix 5A. Additional Detail on 2006 Ecological Screening Assessment .................... 5A-1 

6 Appendix 6A. Additional Detail on 2006 Ecological Screening Assessment .................... 6A-1 
vi

1 LIST OF FIGURES 

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

3 Figure 1-2. Overview of approach to reviewing the secondary standard. ............................... 1-34

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

5 2010 period....................................................................................2-2

Figure 2-2. Trend in U.S. annual 4th 6 highest daily maximum 8-hour O3 concentrations in ppb, 

7 2000 to 2012. .......................................................................................................... 2-4

8 Figure 2-3. Map of 8-hour O3 design values in ppb for the 2009-2011 period. ........................ 2-4

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

10 Figure 2-5. Trend in U.S. annual W126 concentrations in ppm-hrs, 2000 to 2012. ................. 2-6

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

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

13 Figure 2-8. Map of 2007 CMAQ-estimated seasonal mean natural background O3 levels from 

14 zero-out modeling. ................................................................................................ 2-14

15 Figure 2-9. Map of 2007 CMAQ-estimated seasonal mean North American background O3

16 levels from zero-out modeling. ............................................................................. 2-15

17 Figure 2-10. Map of 2007 CMAQ-estimated seasonal mean United States background O3 levels 

18 from zero-out modeling. ....................................................................................... 2-15

19 Figure 2-11. Map of U.S. background percent contribution to seasonal mean O3 based on 2007 

20 CMAQ zero-out modeling. ................................................................................... 2-18

21 Figure 2-12. Map of apportionment-based U.S. background percent contribution to seasonal 

22 mean O3 based on 2007 CAMx source apportionment modeling......................... 2-18

23 Figure 2-13. Distributions of absolute estimates of apportionment-based U.S. Background (all 

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

25 simulation. ............................................................................................................. 2-19

26 Figure 2-14. Distributions of the relative proportion of apportionment-based U.S. Background to 

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

28 apportionment 

29 simulation....................................................................................2-19

30 Figure 2-15. Fractional influence of background sources to W126 levels in four sample 

31 locations. Model estimates based on 2007 CMAQ zero-out modeling. .............. 2-23

32 Figure 2-16. Differences in seasonal mean O3 between the NAB and NB scenarios. ............... 2-25

33 Figure 2-17. Percent contribution of U.S. anthropogenic emissions to total seasonal mean MDA8 

34 O3 levels, based on 2007 source apportionment modeling...........................2-26

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

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

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

38 department visits in studies that presented all-year and/or seasonal results. ........ 3-35
vii

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

2 mortality. ............................................................................................................... 3-49

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

4 distribution of area-wide averaged O3 

5 concentrations...............................................................................3-67

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

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

8 concentrations...... ................................................................................................ 3-69

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

10 ambient O3 concentration study metric. ................................................................ 3-77

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

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

13 (averaged over 2006 to 2010). .............................................................................. 3-90

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

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

16 (worst-case year, 2006 to 2010). ........................................................................... 3-91

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

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

19 (Averaged over 2006 to 2010) .............................................................................. 3-92

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

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

22 (Worst-Case Year, 2006 to 2010) ......................................................................... 3-93

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

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

25 the current standard (Averaged over 2006 to 2010) ........................................... 3-100

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

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

28 the current standard (Worst-Case Year from 2006 to 2010) .............................. 3-101

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

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

31 the current standard (Averaged over 2006 to 2010) ........................................... 3-102

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

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

34 meet the current standard (Worst-Case Year from 2006 to 2010) ...................... 3-103

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

36 meet the current standard. ................................................................................... 3-109

37 Figure 3-16. Estimated O3-associated mortality attributable to days above various area-wide 

38 average O3 concentrations, with air quality adjusted to just meet current standard 

39 (2007 Model Adjustment Year). ......................................................................... 3-110
viii

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

2 O3..........................................................................................3-111

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

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

5 alternative standards (averaged over 2006 to 2010) ............................................. 4-21

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

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

8 alternative standards (worst-case year from 2006 to 2010) .................................. 4-22

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

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

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

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

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

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

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

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

17 current and potential alternative standards (averaged over 2006 to 2010) ........... 4-28

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

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

20 current and potential alternative standards (worst-case year from 2006 to 2010) 4-29

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

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

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

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

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

26 current and potential alternative standards (worst-case year from 2006 to 2010) 4-31

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

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

29 standard levels) – Total Risk ................................................................................. 4-36

30 Figure 4-10. Estimates of Total Mortality Attributable to Days with 8-Hour Area-Wide O3

31 Concentrations at or above 20, 40, or 60 ppb, Summed Across Urban Case Study 

32 Areas (Air Quality Adjusted to Current and Potential alternative standard levels)

33 .............................................................................................4-37

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

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

36 Potential alternative standard levels) – Total Risk ............................................... 4-38

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

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

39 standard levels) – Total Risk ................................................................................. 4-39
ix

1 Figure 4-13. Estimates of Total Mortality Attributable to Days with 8-Hour Area-Wide O3

2 Concentrations at or above 20, 40, or 60 ppb - Risks Summed Across Urban Case 

3 Study Areas and Expressed Relative to 75 ppb........................................4-47

4 Figure 5-1. Relative biomass loss in seedlings for 12 studied species in response to seasonal 

5 ozone concentrations in terms of seasonal W126 index values. ........................... 5-14

6 Figure 5-2. Relationship of tree seedling percent biomass loss with seasonal W126 index. . 5-15

7 Figure 5-3. Association of O3 with cottonwood biomass in gradient downwind of New York 

8 City..........................................................................................5-17

9 Figure 5-4. Relative biomass loss of Ponderosa Pine for air quality model-adjusted to just meet 

10 the current standard. .............................................................................................. 5-24

11 Figure 5-5. Cumulative proportion of sites with a) any foliar injury or b) elevated injury 

12 present, by moisture category. .............................................................................. 5-48
x

1 LIST OF TABLES 

2 Table 2-1. Seasonal mean MDA8 O3 (ppb), seasonal mean apportionment-based USB 

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

4 (all site-days) in the 12 REA urban case study areas (%). .................................... 2-21

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

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

7 (site-days > 60 ppb) in the 12 REA urban study areas (%). ................................. 2-21

8 Table 2-3. Fractional contribution of apportionment-based USB in the 12 REA urban study 

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

10 fractions of seasonal means). ................................................................................ 2-21

11 Table 2-4. Seasonal mean MDA8 O3 (ppb), seasonal mean USB contribution (ppb), and 

12 fractional USB contribution to total O3 (all site-days) in the 12 REA urban case 

13 study areas 

14 (%)............................................................................................2-22

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

16 exposures to ozone concentrations below 75 ppb in young, healthy adults. ........ 3-59

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

18 concentrations below 75 ppb. ............................................................................... 3-61

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

20 locations that would have met the current standard during study periods. ........... 3-64

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

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

23 evaluated by Silverman and Ito, 2010) ................................................................. 3-68

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

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

26 study area evaluated by Strickland et al., 2010) ................................................... 3-70

Table 3-6. Number of study cities with 4th 27 highest 8-hour daily maximum concentrations 

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

29 ................................................................................................3-75

30 Table 3-7. Prevalence of asthma by age in the U.S................................................3-82

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

32 standards with levels of 70, 65, and 60 ppb .......................................................... 4-13

Table 4-2. Number of study cities with 3-year averages of 4th 33 highest 8-hour daily max 

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

35 presented in Bell et al. (2006) ............................................................................... 4-15

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

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

38 potential alternative standards. ............................................................................. 4-18

39 Table 4-4. Summary of Estimated Exposures of Concern for Potential alternative standard 

40 levels of 70, 65, and 60 ppb in Urban Case Study Areas ..................................... 4-45
xi

1 Table 4-5. Summary of Estimated Lung Function Decrements for Potential alternative 

2 standard levels of 70, 65, and 60 ppb in Urban Case Study 

3 Areas.........................................................................................4-46

4 Table 5-1. Ozone concentrations associated with effects on tulip poplar in southern 

5 Appalachian Mountains (2001-2003). .................................................................. 5-16

6 Table 5-2. Examples of counties containing Class I areas where recent air quality might be 

7 expected to meet the current standard and where 3-yr W126 index values are above 

8 15 ppm-hrs. ........................................................................................................... 5-19

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

10 and carbon storage. ............................................................................................... 5-20

11 Table 5-4. Summary of methodology by which national surface of average W126 index values 

12 was derived for each air quality scenario.............................................................. 5-21

13 Table 5-5. Exposure, risk and ecosystem services analyses related to crop yield. ................ 5-33

14 Table 5-6. Visible foliar injury incidence in four National Wildlife Refuges. ...................... 5-41

15 Table 5-7. Exposure, risk and ecosystem services analyses related to visible foliar injury. .. 5-45

16 Table 5-8. Benchmark criteria for O3 exposure and relative soil moisture used in screening17 level assessment of parks ..................................................................5-50

18 Table 6-1. Seedling and crop growth or yield reductions estimated for O3 exposure over a 

19 season. ................................................................................................................... 6-19

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

21 in WREA air quality scenarios. ............................................................................ 6-24

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

23 greater than 2%. .................................................................................................... 6-25

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

25 important crops. .................................................................................................... 6-26

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

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

28 areas..........................................................................................6-28
ES-1

1 EXECUTIVE SUMMARY 

2

3 This second draft Policy Assessment (PA) has been prepared by staff in the 

4 Environmental Protection Agency’s (EPA) Office of Air Quality Planning and Standards 

5 (OAQPS) as part of the Agency’s ongoing review of the primary (health-based) and secondary 

6 (welfare-based) national ambient air quality standards (NAAQS) for ozone (O3). It presents 

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

8 and technical information that informs this review. Preliminary staff conclusions are presented 

9 regarding the adequacy of the current standards and, as appropriate, potential alternative 

10 standards appropriate for consideration in this review. Staff analyses in this second draft PA are 

11 based on the scientific and technical information, as well as uncertainties and limitations related 

12 to this information, assessed in other EPA documents, including the scientific assessment 

13 presented in the Integrated Science Assessment for Ozone, the second draft Health Risk and 

14 Exposure Assessment for Ozone and the second draft Welfare Risk and Exposure Assessment for 

15 Ozone. The final PA is intended to “bridge the gap” between the relevant scientific evidence and 

16 technical information and the judgments required of the EPA Administrator in determining 

17 whether to retain or revise the current standards. Development of the PA is also intended to 

18 facilitate advice and recommendations on the standards to the Administrator from an 

19 independent scientific review committee, the Clean Air Scientific Advisory Committee 

20 (CASAC), as provided for in the Clean Air Act (CAA). 

21 The overarching questions in this review, as in all NAAQS reviews, regard the support 

22 provided by the currently available scientific evidence and exposure/risk-based information for 

23 the adequacy of the current standards and the extent to which the scientific evidence and 

24 technical information provides support for concluding that consideration of alternative standards 

25 may be appropriate. Comments and recommendations from CASAC and public comments based 

26 on review of this draft PA will inform final staff conclusions and the presentation of information 

27 in the final PA.

28 Health Effects and Review of the Primary Standard

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

30 review, documents the effects of O3 in ambient air on health. In particular, O3 affects the 

31 respiratory system, posing greatest hazard to those with respiratory disease and those with 

32 highest exposures, including children with asthma. The evidence indicates that higher exposures 

33 and repeated occurrence of exposures lead to more severe effects, including increased 

34 susceptibility to other respiratory stressors, and that higher exposures lead to greater prevalence 

35 of effects among the exposed population. Based on the staff evaluation presented in this draft 
ES-2

1 document, staff preliminarily concludes that the currently available evidence and exposure and

2 risk information call into question the adequacy of the current primary standard and that 

3 consideration should be given to revising the standard to provide increased public health 

4 protection. With regard to potential alternative standards, staff concludes it is appropriate to 

5 consider standards with the same indicator, averaging time and form as the current standard with 

6 alternative levels within the range from 70 ppb to 60 ppb.

7 In drawing these preliminary conclusions, staff additionally notes that the final decision 

8 on the adequacy of the current standard and consideration of potential alternative standards is 

9 largely a public health policy judgment to be made by the Administrator, drawing upon the 

10 scientific information as well as judgments about how to consider the range and magnitude of 

11 uncertainties that are inherent in the scientific evidence and technical analyses. 

12 Welfare Effects and Review of the Secondary Standard

13 The longstanding evidence base, strengthened since the last review, documents the 

14 welfare-related effects of O3 in ambient air. In particular, O3 affects vegetation and poses risk of 

15 related effects on terrestrial ecosystems. Based on the staff evaluation presented in this draft 

16 document, staff preliminarily concludes that the currently available evidence and exposure and 

17 risk information call into question the adequacy of the current secondary standard and that 

18 consideration should be given to revising the standard to provide increased public welfare 

19 protection. In considering the level of protection achieved by potential alternative standards, 

20 staff preliminarily concludes it is appropriate for the Administrator to judge O3 welfare impacts 

21 using the W126-based cumulative seasonal index, defined as an index of the sum of weighted 

22 hourly concentrations, cumulated over 12 hours per day (8 am to 8 pm) during the consecutive 

23 three-month period within the O3 season with the maximum index value. With regard to 

24 potential alternative standards, staff preliminarily concludes it is appropriate to consider 

25 standards in terms of the W126-based cumulative seasonal metric with a form averaged across 

26 three consecutive years and levels extending somewhat above 15 ppm-hrs (e.g., to 17 ppm-hrs) 

27 down to 7 ppm-hrs. 

28 In drawing these preliminary conclusions, staff additionally notes that the final decision 

29 on the adequacy of the current standard and consideration of potential alternative standards is 

30 largely a public welfare policy judgment to be made by the Administrator, drawing upon the 

31 scientific information as well as judgments about how to consider the range and magnitude of 

32 uncertainties that are inherent in the scientific evidence and technical analyses. 
1-1

1 1 INTRODUCTION 

2 1.1 PURPOSE 

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

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

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

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

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

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

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

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

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

information in the ISA and REAs for EPA’s consideration.112 The PA provides a transparent 

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

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

15 appropriate to consider. 

16 When final, the PA is intended to help “bridge the gap” between the Agency’s scientific 

17 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

18 In 

19 evaluating the adequacy of the current standard and whether it is appropriate to consider 

20 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,421 and level. These 

22 elements, which together serve to define each standard, must be considered collectively in 

23 evaluating the health and welfare protection afforded by the O3 standards. The PA integrates and 

24 interprets the information from the ISA and REAs to frame policy options for consideration by 

25 the Administrator. In so doing, the PA recognizes that the selection of a specific approach to 

26 reaching final decisions on the primary and secondary O3 standards will reflect the judgments of 

27 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

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

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

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

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

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

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

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

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

9 The decision whether to prepare one or more drafts of the PA is influenced by 

10 preliminary staff conclusions and associated CASAC advice and public comment, among other 

11 factors. Typically, as in this review, staff prepares a second draft PA where the available 

12 information calls into question the adequacy of the current standard and analyses of potential 

13 alternative standards are developed taking into consideration CASAC advice and public 

14 comment. In such cases, a second draft PA includes preliminary staff conclusions regarding 

15 potential alternative standards and undergoes CASAC review and public comment prior to 

preparation of the final PA.516

17 In this second draft of the PA for the review of the O3 NAAQS, we consider the scientific 

18 and technical information available in this review as assessed in the Integrated Science 

19 Assessment for O3 and Related Photochemical Oxidants (ISA, U.S. EPA, 2013), prepared by 

20 EPA’s National Center for Environmental Assessment (NCEA), and the second drafts of the 

21 quantitative human exposure and health risk assessment and welfare risk assessment documents 

22 (HREA, U.S. EPA, 2014a; WREA, U.S. EPA, 2014b). The evaluation and preliminary staff 

23 conclusions presented in this second draft PA have been informed by comments and advice 

24 received from CASAC in their review of the first draft PA and of the other draft Agency 

25 documents prepared for this NAAQS review. Review and comments on this second draft PA 

26 from CASAC and the public will inform the final evaluation and staff conclusions in the final 

27 PA. 

28 Beyond informing the EPA Administrator and facilitating the advice and 

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

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

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

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

 

5

 When such analyses are not undertaken, a second draft PA may not be warranted. 
1-3

1 1.2 BACKGROUND 

2 1.2.1 Legislative Requirements 

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

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

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

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

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

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

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

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

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

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

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

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

15 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

16 

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

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

19 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

20

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

22 intended to address uncertainties associated with inconclusive scientific and technical 

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

24 degree of protection against hazards that research has not yet identified. See Lead Industries 

25 Association v. EPA, 647 F.2d 1130, 1154 (D.C. Cir 1980); American Petroleum Institute v. 

26 Costle, 665 F.2d 1176, 1186 (D.C. Cir. 1981); American Farm Bureau Federation v. EPA, 559 F. 

27 3d 512, 533 (D.C. Cir. 2009); Association of Battery Recyclers v. EPA, 604 F. 3d 613, 617-18 

28 (D.C. Cir. 2010). Both kinds of uncertainties are components of the risk associated with pollution 

29 at levels below those at which human health effects can be said to occur with reasonable 

 

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

1 scientific certainty. Thus, in selecting primary standards that provide an adequate margin of 

2 safety, the Administrator is seeking not only to prevent pollution levels that have been 

3 demonstrated to be harmful but also to prevent lower pollutant levels that may pose an 

4 unacceptable risk of harm, even if the risk is not precisely identified as to nature or degree. The 

5 CAA does not require the Administrator to establish a primary NAAQS at a zero-risk level or at 

6 background concentration levels, see Lead Industries v. EPA, 647 F.2d at 1156 n.51; State of 

7 Mississippi v. EPA, 723 F. 3d 246, 255, 262-63 (D.C. Cir. 2013), but rather at a level that 

8 reduces risk sufficiently so as to protect public health with an adequate margin of safety. 

9 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)810 at risk, 

11 and the kind and degree of the uncertainties that must be addressed. The selection of any 

12 particular approach for providing an adequate margin of safety is a policy choice left specifically 

13 to the Administrator’s judgment. See Lead Industries Association v. EPA, 647 F.2d at 1161-62; 

14 State of Mississippi, 723 F. 3d at 265. 

15 In setting primary and secondary standards that are “requisite” to protect public health 

16 and welfare, respectively, as provided in section 109(b), EPA’s task is to establish standards that 

17 are neither more nor less stringent than necessary for these purposes. In so doing, the EPA may 

18 not consider the costs of implementing the standards. See generally, Whitman v. American 

19 Trucking Associations, 531 U.S. 457, 465-472, 475-76 (2001). Likewise, “[a]ttainability and 

20 technological feasibility are not relevant considerations in the promulgation of national ambient 

21 air quality standards.” American Petroleum Institute v. Costle, 665 F. 2d at 1185. 

22 Section 109(d)(1) requires that “not later than December 31, 1980, and at 5-year intervals 

23 thereafter, the Administrator shall complete a thorough review of the criteria published under 

24 section 108 and the national ambient air quality standards . . . and shall make such revisions in 

25 such criteria and standards and promulgate such new standards as may be appropriate . . . .” 

26 Section 109(d)(2) requires that an independent scientific review committee “shall complete a 

27 review of the criteria . . . and the national primary and secondary ambient air quality standards . . 

28 . and shall recommend to the Administrator any new . . . standards and revisions of existing 

29 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. 930

 

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. 

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. 
1-5

1 1.2.2 History of O3 NAAQS Reviews 

2 Table 1-1 summarizes the O3 NAAQS that the EPA has promulgated to date. In each 

3 review, the EPA set the secondary standard at a level identical to the primary standard. These 

4 reviews are briefly described below. 

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

6 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

7

8 The EPA first established primary and secondary NAAQS for photochemical oxidants in 

9 1971 (36 FR 8186, April 30, 1971). The EPA set both primary and secondary standards at a level 

10 of 0.08 parts per million (ppm), 1-hr average, total photochemical oxidants, not to be exceeded 

11 more than one hour per year. The EPA based the standards on scientific information contained in 

12 the 1970 Air Quality Criteria for Photochemical Oxidants (U.S. DHEW, 1970). We initiated the 

13 first periodic review of the NAAQS for photochemical oxidants in 1977. Based on the 1978 Air 

14 Quality Criteria for Ozone and Other Photochemical Oxidants (U.S. EPA, 1978), the EPA 

15 published proposed revisions to the original NAAQS in 1978 (43 FR 16962) and final revisions 

16 in 1979 (44 FR 8202). At that time, the EPA revised the level of the primary and secondary 

17 standards from 0.08 to 0.12 ppm and changed the indicator from photochemical oxidants to O3, 

18 and the form of the standards from a deterministic to a statistical form. This statistical form 
1-6

1 defined attainment of the standards as occurring when the expected number of days per calendar 

2 year with maximum hourly average concentration greater than 0.12 ppm equaled one or less. 

3 Following the final decision in the 1979 review, the City of Houston challenged the 

4 Administrator’s decision arguing that the standard was arbitrary and capricious because natural 

5 O3 concentrations and other physical phenomena in the Houston area made the standard 

6 unattainable in that area. The U.S. Court of Appeals for the District of Columbia circuit (D.C. 

7 Circuit) rejected this argument, stating (as noted above) that attainability and technological 

8 feasibility are not relevant considerations in the promulgation of the NAAQS. The court also 

9 noted that the EPA need not tailor the NAAQS to fit each region or locale, pointing out that 

10 Congress was aware of the difficulty in meeting standards in some locations and had addressed 

11 this difficulty through various compliance related provisions in the Act. See API v. Costle, 665 

12 F.2d 1176, 1184-6 (D.C. Cir. 1981). In 1982, we announced plans to revise the 1978 Air Quality 

13 Criteria document (47 FR 11561), and in 1983, we initiated the second periodic review of the O3

14 NAAQS (48 FR 38009). We subsequently published the 1986 Air Quality Criteria for Ozone 

15 and Other Photochemical Oxidants (U.S. EPA, 1986) and the 1989 Staff Paper (U.S. EPA, 

16 1989). Following publication of the 1986 Air Quality Criteria Document (AQCD), a number of 

17 scientific abstracts and articles were published that appeared to be of sufficient importance 

18 concerning potential health and welfare effects of O3 to warrant preparation of a Supplement 

19 (U.S. EPA, 1992). On August 10, 1992, under the terms of a court order, the EPA published a 

20 proposed decision to retain the existing primary and secondary standards. (57 FR 35542). The 

21 notice explained that the proposed decision would complete EPA’s review of information on 

22 health and welfare effects of O3 assembled over a 7-year period and contained in the 1986 

23 AQCD and its 1992 Supplement. The proposal also announced EPA’s intention to proceed as 

24 rapidly as possible with the next review of the air quality criteria and standards for O3 in light of 

25 emerging evidence of health effects related to 6- to 8-hour O3 exposures. On March 9, 1993, the 

26 EPA concluded the review by affirming its proposed decision to retain the existing primary and 

27 secondary standards. (58 FR 13008). 

28 In August 1992, we announced plans to initiate the third periodic review of the air quality 

29 criteria and O3 NAAQS (57 FR 35542). In December 1996, the EPA proposed to replace the then 

30 existing 1-hour primary and secondary standards with 8-hour average O3 standards set at a level 

31 of 0.08 ppm (equivalent to 0.084 ppm using standard rounding conventions) (61 FR 65716). The 

32 EPA also proposed to establish a new distinct secondary standard using a biologically-based 

33 cumulative, seasonal form. The EPA completed this review on July 18, 1997 (62 FR 38856) by 

34 setting the primary standard at a level of 0.08 ppm, based on the annual fourth-highest daily 

35 maximum 8-hr average concentration, averaged over three years, and setting the secondary 

36 standard identical to the revised primary standard. In reaching this decision, the EPA identified 
1-7

1 several reasons supporting its decision to reject a potential alternate standard set at 0.07 ppm. 

2 Most importantly, the EPA pointed out the scientific uncertainty at lower concentrations and 

3 placed significant weight on the fact that no CASAC panel member supported a standard level 

4 set lower than 0.08 ppm (62 FR 38868). In addition to noting the uncertainties in the health 

5 evidence for exposure concentrations below 0.08 ppm and the advice of CASAC, the EPA noted 

6 that a standard set at a level of 0.07 ppm would be closer to peak background concentrations that 

7 infrequently occur in some areas due to nonanthropogenic sources of O3 precursors (62 FR 

8 38856, 38868; July 18, 1997). 

9 On May 14, 1999, in response to challenges by industry and others to EPA’s 1997 

10 decision, the U.S. Court of Appeals for the District of Columbia Circuit remanded the O3

11 NAAQS to the EPA, finding that section 109 of the Act, as interpreted by the EPA, effected an 

12 unconstitutional delegation of legislative authority. American Trucking Assoc. vs. EPA, 175 

13 F.3d 1027, 1034-1040(D.C. Cir. 1999) (“ATA I”). In addition, the court directed that, in 

14 responding to the remand, the EPA should consider the potential beneficial health effects of O3

15 pollution in shielding the public from the effects of solar ultraviolet (UV) radiation, as well as 

16 adverse health effects. Id. At 1051-53. In 1999, the EPA petitioned for rehearing en banc on 

17 several issues related to that decision. The court granted the request for rehearing in part and 

18 denied it in part, but declined to review its ruling with regard to the potential beneficial effects of 

19 O3 pollution. 195 F3d 4, 10 (D.C Cir., 1999) (“ATA II”). On January 27, 2000, the EPA 

20 petitioned the U.S. Supreme Court for certiorari on the constitutional issue (and two other 

21 issues), but did not request review of the ruling regarding the potential beneficial health effects 

22 of O3. On February 27, 2001, the U.S. Supreme Court unanimously reversed the judgment of the 

23 D.C. Circuit on the constitutional issue. Whitman v. American Trucking Assoc., 531 U. S. 457, 

24 472-74 (2001) (holding that section 109 of the CAA does not delegate legislative power to the 

25 EPA in contravention of the Constitution). The Court remanded the case to the D.C. Circuit to 

26 consider challenges to the O3 NAAQS that had not been addressed by that court’s earlier 

27 decisions. On March 26, 2002, the D.C. Circuit issued its final decision on remand, finding the 

28 1997 O3 NAAQS to be “neither arbitrary nor capricious,” and so denying the remaining petitions 

29 for review. American Trucking Associations, Inc. v EPA, 283 F.3d 355, 379 (D.C Cir., 

30 2002)(“ATA III”). 

31 Specifically, in ATA III, the D.C. Circuit upheld EPA’s decision on the 1997 O3 standard 

32 as the product of reasoned decision-making. The Court made clear that the most important 

33 support for EPA’s decision was the health evidence and the concerns it raised about setting a 

34 standard level below 0.08 ppm. (“the record is replete with references to studies demonstrating 

35 the inadequacies of the old one-hour standard”, as well as extensive information supporting the 

36 change to an 8-hour averaging time). 283 F 3d at 378. The Court also pointed to the significant 
1-8

1 weight that the EPA properly placed on the advice it received from CASAC. Id. at 379. The 

2 court further noted that “although relative proximity to peak background ozone concentrations 

3 did not, in itself, necessitate a level of 0.08, EPA could consider that factor when choosing 

4 among the three alternative levels.” Id. 

5 Independently of the litigation, the EPA also responded to the Court’s remand to 

6 consider the potential beneficial health effects of O3 pollution in shielding the public from effects 

7 of solar (ultraviolet or UV-B) radiation. The EPA provisionally determined that that the 

8 information linking changes in patterns of ground-level O3 concentrations to changes in relevant 

9 patterns of exposures to ultraviolet (UV-B) radiation of concern to public health was too 

10 uncertain, at that time, to warrant any relaxation in 1997 O3 NAAQS. The EPA also expressed 

11 the view that any plausible changes in UV-B radiation exposures from changes in patterns of 

12 ground-level O3 concentrations would likely be very small from a public health perspective. In 

13 view of these findings, the EPA proposed to leave the 1997 8-hour NAAQS unchanged (66 FR 

14 57268, Nov. 14, 2001). After considering public comment on the proposed decision, the EPA 

15 published its final response to this remand on January 6, 2003, re-affirming the 8-hour O3

16 NAAQS set in 1997 (68 FR 614). 

17 The EPA initiated the fourth periodic review of the air quality criteria and O3 standards in 

18 September 2000 with a call for information (65 FR 57810). The schedule for completion of that 

19 review was ultimately governed by a consent decree resolving a lawsuit filed in March 2003 by 

20 plaintiffs representing national environmental and public health organizations, who maintained 

21 that EPA was in breach of a mandatory legal duty to complete review of the O3 NAAQS within a 

22 statutorily-mandated deadline. On July 11, 2007, the EPA proposed to revise the level of the 

23 primary standard within a range of 0.075 to 0.070 ppm. (72 FR 37818). Documents supporting 

24 this proposed decision included the Air Quality Criteria for Ozone and Other Photochemical 

25 Oxidants (U.S. EPA, 2006) and the Staff Paper (U.S EPA, 2007a) and related technical support 

26 documents. The EPA also proposed two options for revising the secondary standard: (1) replace 

27 the current standard with a cumulative, seasonal standard, expressed as an index of the annual 

28 sum of weighted hourly concentrations cumulated over 12 daylight hours during the consecutive 

29 3-month period within the O3 season with the maximum index value, set at a level within the 

30 range of 7 to 21 ppm-hrs, and (2) set the secondary standard identical to the proposed primary 

31 standard. The EPA completed the review with publication of a final decision on March 27, 2008 

32 (73 FR 16436). In that final rule, the EPA revised the NAAQS by lowering the level of the 8-

33 hour primary O3 standard from 0.08 ppm to 0.075 ppm, not otherwise revising the primary 

34 standard, and adopting a secondary standard identical to the revised primary standard. In May 

35 2008, state, public health, environmental, and industry petitioners filed suit challenging EPA’s 

36 final decision on the 2008 O3 standards. On September 16, 2009, the EPA announced its 
1-9

1 intention to reconsider the 2008 O3 standards, and initiated a rulemaking to do so. At EPA’s 

2 request, the Court held the consolidated cases in abeyance pending EPA’s reconsideration of the 

3 2008 decision. 

4 On January 19, 2010 (75 FR 2938), the EPA issued a notice of proposed rulemaking to 

5 reconsider the 2008 final decision. In that notice, the EPA proposed that further revisions of the 

6 primary and secondary standards were necessary to provide a requisite level of protection to 

7 public health and welfare. The EPA proposed to decrease the level of the 2008 8-hour primary 

8 standard from 0.075 ppm to a level within the range of 0.060 to 0.070 ppm, and to change the 

9 secondary standard to a new cumulative, seasonal standard expressed as an annual index of the 

10 sum of weighted hourly concentrations, cumulated over 12 hours per day (8 am to 8 pm), during 

11 the consecutive 3-month period within the O3 season, with a maximum index value set at a level 

12 within the range of 7 to 15 ppm-hours. The Agency also solicited CASAC review of the 

13 proposed rule on January 25, 2010 and solicited additional CASAC advice on January 26, 2011. 

14 After considering comments from CASAC and the public, the EPA prepared a draft final rule, 

15 which was submitted for interagency review pursuant to Executive Order 12866. On September 

16 2, 2011, consistent with the direction of the President, the Administrator of the Office of 

17 Information and Regulatory Affairs (“OIRA”), Office of Management and Budget (“OMB”), 

18 returned the draft final rule to the EPA for further consideration. In view of this return and the 

19 timing of the Agency’s ongoing periodic review of the O3 NAAQS required under Clean Air Act 

20 section 109 (as announced on September 29, 2008), the EPA decided to coordinate further 

21 proceedings on its voluntary rulemaking on reconsideration with that ongoing periodic review, 

22 by deferring the completion of its voluntary rulemaking on reconsideration until it completes its 

23 statutorily-required periodic review. 

24 In light of EPA’s decision to consolidate the reconsideration with the current review, the 

25 Court proceeded with the litigation on the 2008 final decision. On July 23, 2013, the D.C. Circuit 

26 Court of Appeals upheld EPA’s 2008 primary O3 standard, but remanded the 2008 secondary 

27 standard to the EPA. State of Mississippi v. EPA. 723 F. 3d 246. With respect to the primary 

28 standard, the court first held that the EPA reasonably determined that the existing standard was 

29 not requisite to protect public health with an adequate margin of safety, and consequently 

30 required revision. Specifically, the court noted that there were “numerous epidemiological 

31 studies linking health effects to exposure to ozone levels below 0.08 ppm and clinical human 

32 exposure studies finding a causal relationship between health effects and exposure to ozone 

33 levels at and below 0.08 ppm”. Id. at 257. The court also specifically endorsed the weight of 

34 evidence approach utilized by EPA in its deliberations. Id. at 256. 

35 The court went on to reject arguments that EPA should have adopted a more stringent 

36 primary standard. Dismissing arguments that a single clinical study (properly interpreted by 
1-10

1 EPA) to show effects at 0.06 ppm necessitated a standard level lower than that selected, the court 

2 noted that this was a single, limited study. Id. at 262. With respect to the epidemiologic evidence, 

3 the court accepted EPA’s argument that there could be legitimate uncertainty that a causal 

4 relationship between O3 and 8-hour exposures less than 0.075 ppm exists, so that associations at 

5 lower levels reported in epidemiologic studies did not necessitate a more stringent standard. Id.

at 264-65.10 6

7 The court also rejected arguments that an 8-hour primary standard of 0.075 ppm failed to 

8 provide an adequate margin of safety, noting that margin of safety considerations involved policy 

9 judgments by the agency, and that by setting a standard “appreciably below” the level of the 

10 current standard (0.08 ppm), the agency had made a reasonable policy choice . Id. Finally, the 

11 court rejected arguments that EPA’s decision was inconsistent with CASAC’s scientific 

12 recommendations because CASAC had been insufficiently clear in its recommendations whether 

13 it was providing scientific or policy recommendations, and EPA had reasonably addressed 

14 CASAC’s policy recommendations. Id. at 269-70. 

15 With respect to the secondary standard, the court held that because EPA had failed to 

16 identify a level of air quality requisite to protect public welfare, EPA’s comparison between the 

17 primary and secondary standards for determining if requisite protection for public welfare was 

18 afforded by the primary standard was inherently arbitrary. The court thus rejected EPA’s 

19 determination that the revised 8-hour primary standard afforded requisite protection of public 

20 welfare, and remanded the standard to EPA. Id. at 272-73.

21 1.2.3 Current O3 NAAQS Review 

22 On September 29, 2008, the EPA announced the initiation of a new periodic review of 

23 the air quality criteria for O3 and related photochemical oxidants and issued a call for 

24 information in the Federal Register (73 FR 56581, Sept. 29, 2008). A wide range of external 

25 experts, as well as EPA staff, representing a variety of areas of expertise (e.g., epidemiology, 

26 human and animal toxicology, statistics, risk/exposure analysis, atmospheric science, ecology, 

27 biology, plant science, ecosystem services) participated in a workshop. This workshop was held 

28 on October 28-29, 2008 in Research Triangle Park, NC. The workshop provided an opportunity 

29 for a public discussion of the key policy-relevant issues around which the EPA would structure 

30 this O3 NAAQS review and the most meaningful new science that would be available to inform 

31 our understanding of these issues. 

 

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 262, 269 (internal citations omitted). 
1-11

1 Based in part on the workshop discussions, the EPA developed a draft IRP outlining the 

2 schedule, process, and key policy-relevant questions that would guide the evaluation of the air 

3 quality criteria for O3 and the review of the primary and secondary O3 NAAQS. A draft of the 

4 integrated review plan was released for public review and comment in September 2009. This IRP 

5 was the subject of a consultation with the CASAC on November 13, 2009 (74 FR 54562; 

October 22, 2009).11 6 We considered comments received from that consultation and from the 

7 public in finalizing the plan and in beginning the review of the air quality criteria. The EPA’s 

8 overall plan and schedule for this review is presented in the Integrated Review Plan for the 

Ozone National Ambient Air Quality Standards.

12 9

10 As part of the process of preparing the O3 ISA, NCEA hosted a peer review workshop in 

11 October 29-30, 2008 (73 FR 56581, September 29, 2008) on preliminary drafts of key ISA 

12 chapters. The CASAC and the public reviewed the first external review draft ISA (U.S. EPA, 

13 2011b; 76 FR 10893, February 28, 2011) at a meeting held in May 19-20, 2011 (76 FR 23809; 

14 April 28, 2011). Based on CASAC and public comments, NCEA prepared a second draft ISA 

15 (U.S. EPA, 2011c; 76 FR 60820, September 30, 2011). CASAC and the public reviewed this 

16 draft at a January 9-10, 2012 (76 FR 236, December 8, 2011) meeting. Based on CASAC and 

17 public comments, NCEA prepared a third draft ISA (U.S. EPA 2012a; 77 FR 36534; June 19, 

18 2012), which was reviewed at a CASAC meeting in September 2012. The final ISA was released 

19 in February 2013. 

20 The EPA presented its plans for conducting the Risk and Exposure Assessments (REAs) 

21 that build on the scientific evidence presented in the ISA, in two planning documents titled 

22 Ozone National Ambient Air Quality Standards: Scope and Methods Plan for Health Risk and 

23 Exposure Assessment and Ozone National Ambient Air Quality Standards: Scope and Methods 

Plan for Welfare Risk and Exposure Assessment (henceforth, Scope and Methods Plans).13 24 

25 These planning documents outlined the scope and approaches that staff planned to use in 

26 conducting quantitative assessments, as well as, key issues that would be addressed as part of the 

27 assessments. We released these documents for public comment in April 2011, and consulted with 

28 CASAC on May 19-20, 2011 (76 FR 23809; April 28, 2011). In designing and conducting the 

29 initial health risk and welfare risk assessments, we considered CASAC comments (Samet 2011) 

30 on the Scope and Methods Plans and also considered public comments. In May 2012, we issued 

 

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 

13 EPA-452/P-11-001 and -002; April 2011; Available: 

http://www.epa.gov/ttn/naaqs/standards/ozone/s_o3_2008_pd.html 
1-12

1 a memo titled, Updates to Information Presented in the Scope and Methods Plans for the Ozone 

2 NAAQS Health and Welfare Risk and Exposure Assessments, that described changes to elements 

3 of the scope and methods plans and provided a brief explanation of each change and the reason 

4 for it. 

5 In July 2012, EPA made the first drafts of the Health and Welfare REAs available for 

6 CASAC review and public comment (77 FR 42495, July 19, 3023). The first draft PA was made 

7 available for CASAC review and public comment in August 2012. These documents were 

8 reviewed by CASAC O3 Panel at a public meeting in September 2012. The second draft REAs 

9 and PA have been prepared in consideration of CASAC (Frey and Samet, 2012a, 2012b) and 

10 public comment and will be reviewed by the CASAC O3 Panel at a public meeting in March 

11 2014.

12 1.3 GENERAL APPROACH FOR REVIEW OF THE STANDARDS

13 As described in section 1.1 above, the final PA will present a transparent evaluation and 

14 staff conclusions regarding policy considerations related to reaching judgments about the 

15 adequacy of the current standards and what, if any, revisions may be appropriate to consider. 

16 Preliminary staff considerations and conclusions in this document are based on the available 

17 body of scientific evidence assessed in the ISA (U.S. EPA, 2013), exposure and risk analyses 

18 presented in the 2nd draft REAs (U.S. EPA, 2014a, b), advice and recommendations from 

19 CASAC on the first draft PA and other draft EPA documents in this review, as well as on public 

20 comments. When final, this evaluation and associated conclusions on the range of policy options 

21 that, in staff’s view, could be supported by the available scientific evidence and exposure/risk 

22 information will inform the Administrator’s decisions as to whether the existing primary and/or 

23 secondary O3 standards should be revised and, if so, what revised standard or standards is/are 

24 appropriate. 

25 Staff’s considerations and conclusions related to the current and alternative primary and 

26 secondary O3 standards are framed by a series of key policy-relevant questions, expanding upon 

27 those presented in the IRP at the outset of this review (U.S. EPA, 2011a). Answers to these 

28 questions in the final PA will inform the Administrator’s decisions as to whether, and if so how, 

29 to revise the current O3 standards. The first overarching question is as follows. 

30 Do the currently available scientific evidence and exposure/risk information, as 

31 reflected in the ISA and REAs, support or call into question the adequacy of the 

32 protection afforded by the current O3 standards? 

33 If the answer to this question, which is informed by staff’s consideration of more specific 

34 questions related to the primary and secondary standards, suggests that revision of the current 
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1 standards may be appropriate, then staff further considers the currently available evidence and 

2 information with regard to the following question. 

3 What range of potential alternative standards is appropriate to consider based on 

4 the scientific evidence, air quality analyses, and exposure/risk-based 

5 information? 

6 The general approaches for consideration of these overarching questions in review of the primary 

7 and secondary standards are described separately in sections 1.3.1 and 1.3.2 below. 

8 1.3.1 Approach for the Primary Standard 

9 Staff’s approach in this review of the current primary O3 standard takes into 

10 consideration the approaches used in previous O3 NAAQS reviews. The past and current 

11 approaches described below are both based, fundamentally, on using EPA’s assessment of the 

12 current scientific evidence and associated quantitative analyses to inform the Administrator’s 

13 judgment regarding a primary standard for O3 that is “requisite” (i.e., neither more nor less 

14 stringent than necessary) to protect public health with an adequate margin of safety. 

15 In reaching conclusions on options for the Administrator’s consideration, we note that the 

16 final decision to retain or revise the current primary O3 standard is a public health policy 

17 judgment to be made by the Administrator. This final decision by the Administrator will draw 

18 upon the available scientific evidence for O3-attributable health effects, and on analyses of 

19 population exposures and health risks, including judgments about the appropriate weight to 

20 assign the range of uncertainties inherent in the evidence and analyses. Our general approach to 

21 informing these judgments, discussed more fully below, recognizes that the available health 

22 effects evidence reflects a continuum from relatively higher O3 concentrations, at which 

23 scientists generally agree that health effects are likely to occur, through lower concentrations, at 

24 which the likelihood and magnitude of a response become increasingly uncertain. Therefore, in 

25 developing conclusions in this second draft PA, we are mindful that the Administrator’s ultimate 

26 judgments on the primary standard will most appropriately reflect an interpretation of the 

27 available scientific evidence and exposure/risk information that neither overstates nor understates 

28 the strengths and limitations of that evidence and information. This approach is consistent with 

29 the requirements of sections 108 and 109 of the Act, as well as with how the EPA and the courts 

30 have historically interpreted the Act. 

31 Section 1.3.1.1 below provides an overview of the general approach taken in the last 

32 review of the primary O3 NAAQS (i.e., the 2008 review), and a summary of the rationale for the 

33 decision on the level of the standard in that review (73 FR 16436). Section 1.3.1.2 presents our 

34 approach in the current review, including our approach to considering the health evidence and 
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1 exposure/risk information, and considerations regarding ambient O3 concentrations attributable 

2 to background sources. 

3 1.3.1.1 Approach Used in the Last Review 

4 In the 2008 review of the O3 NAAQS, the Administrator relied upon consideration of the 

5 available scientific evidence and exposure/risk information, the advice and recommendations of 

6 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 7 ). In reaching a decision to revise 

8 the 1997 8-hour primary O3 standard, the Administrator noted that much new evidence had 

9 become available since the 1997 review. He noted that this body of scientific evidence was very 

10 robust and provided consistent and coherent evidence of an array of O3-related respiratory 

11 morbidity effects, and possibly cardiovascular-related morbidity, as well as total nonaccidental 

12 and cardiorespiratory mortality. The Administrator specifically observed that (1) the evidence of 

13 a range of respiratory-related morbidity effects had been considerably strengthened; (2) newly 

14 available evidence from controlled human exposure and epidemiologic studies identified people 

15 with asthma as an important susceptible population for which estimates of respiratory effects in 

16 the general population likely underestimate the magnitude or importance of these effects; (3) 

17 newly available evidence about mechanisms of toxicity more completely explained the 

18 biological plausibility of O3-induced respiratory effects and was beginning to suggest 

19 mechanisms that may link O3 exposure to cardiovascular effects; and (4) there was relatively 

20 strong evidence for associations between short-term O3 concentrations and total nonaccidental 

21 and cardiopulmonary mortality. The Administrator believed that this very robust body of 

22 evidence enhanced our understanding of O3- related effects and provided increased confidence 

23 that various respiratory morbidity effects and other effects marked by indicators of respiratory 

24 morbidity are causally related to O3 exposures, and that the evidence was highly suggestive that 

O3 exposures during the warm O3 season contribute to premature mortality.16 25

26 The Administrator also noted important new health evidence reporting a broad array of 

27 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 28 including people with 

 

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 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. 

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 
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1 asthma or other lung diseases, older adults with increased susceptibility, and those who are likely 

2 to be vulnerable as a result of spending a lot of time outdoors engaged in physical activity (e.g., 

3 especially active children and outdoor workers). 

4 He specifically noted new scientific evidence, which built upon existing evidence, 

5 demonstrating O3-induced lung function effects and respiratory symptoms in some healthy 

6 individuals following exposures down to 80 ppb. He also noted very limited new evidence 

7 demonstrating such effects at exposure concentrations well below 80 ppb. In addition, the 

8 Administrator noted (1) epidemiologic evidence of statistically significant associations with O3-

9 related health effects in areas that likely would have met the then-current standard; (2) 

10 epidemiologic studies conducted in areas that likely would have violated the existing standard 

11 but which nonetheless reported statistically significant associations that generally extended down 

12 to ambient O3 concentrations below the level of that standard; (3) the few studies that had 

13 reported statistically significant associations with respiratory morbidity outcomes and mortality 

14 in subsets of data that included only days with ambient O3 concentrations below the level of the 

15 existing standard; and (4) controlled human exposure studies, together with animal toxicological 

16 studies, that provided considerable support for the biological plausibility of the respiratory 

17 morbidity associations observed in the epidemiologic studies. Based on the available evidence, 

18 the Administrator agreed with the CASAC and the majority of public commenters that the 

19 existing standard was not requisite to protect public health with an adequate margin of safety (FR 

20 73 16471). 

21 Beyond this focus on the available health evidence, the Administrator also considered 

22 estimates of O3 exposures and health risks based on analyses where air quality was adjusted to 

23 simulate just meeting the existing and potential alternative standards. For the various air quality 

24 simulations, he specifically considered the pattern of estimated reductions in O3 exposures across 

25 health benchmark concentrations of 80, 70, and 60 ppb. The 80 ppb benchmark reflected an 

26 exposure concentration for which there was strong evidence for respiratory effects in healthy 

27 people, including airway inflammation, respiratory symptoms, airway hyperresponsiveness, and 

28 impaired lung host defense (U.S. EPA, 2007, section 4.7). The 60 ppb benchmark reflected an 

 

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

1 exposure concentration for which the Administrator judged the evidence of such effects to be 

2 very limited (73 FR 16471). 

3 The Administrator took note of the magnitudes of estimated health risks for a range of 

4 health effects, including moderate and large lung function decrements, respiratory symptoms, 

5 respiratory-related hospital admissions, and nonaccidental and cardiorespiratory mortality. He 

6 recognized that these quantitative risk estimates for a limited number of specific health effects 

7 were indicative of a much broader array of O3-related effects, including various indicators of 

8 morbidity in at-risk populations that we could not analyze in the risk assessment (e.g., school 

9 absences, increased medication use, emergency department visits). The Administrator concluded 

10 that quantitative exposure and risk estimates, as well as the broader array of O3-related health 

11 endpoints that could not be quantified, provided additional support for the evidence-based 

12 conclusion that the existing standard needed to be revised (73 FR 16472). 

13 Based on the above considerations, and consistent with CASAC’s unanimous conclusion 

14 that there was no scientific justification for retaining the existing standard, the Administrator 

15 concluded that the primary O3 standard set in 1997 was not sufficient and thus not requisite to 

16 protect public health with an adequate margin of safety. He further concluded that revision of 

17 this standard was needed to provide increased public health protection (73 FR 16472). 

18 Throughout the 2008 review, CASAC supported a standard level in the range of 60 to 70 

19 ppb (without change to the form, indicator, or averaging time). In a letter to the Administrator on 

20 the second draft Staff Paper, CASAC unanimously recommended “that the current primary 

21 ozone standard be revised and that the level that should be considered for the revised standard be 

22 from 0.060 to 0.070 ppm’’ (60 to 70 ppb) (Henderson, 2006, p. 5). This recommendation, based 

23 in part on the placement of more weight on the evidence for effects following exposures to 60 

24 ppb O3, followed from the CASAC’s more general recommendation that the 1997 standard 

25 needed to be made substantially more protective of human health, particularly for at-risk 

26 populations. In a subsequent letter sent specifically to offer advice to aid the Administrator and 

27 Agency staff in developing the 2007 O3 proposal, CASAC reiterated that Panel members ‘‘were 

28 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 29

30 After considering CASACs comments, the Administrator judged that the appropriate 

31 balance to draw, based on the entire body of evidence and information available in the 2008 

 

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. 723 F. 3d at 269. 
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1 review, was a standard set at a level of 75 ppb (and leaving all other elements of the NAAQS 

2 unchanged). In making this decision, the Administrator placed primary emphasis on the body of 

3 available scientific evidence, while viewing the results of exposure and risk assessments as 

4 providing supporting information. Specifically, the Administrator judged that a standard set at 

5 75 ppb would be appreciably below 80 ppb, the level in controlled human exposure studies at 

6 which adverse effects had been demonstrated at the time, and would provide a significant 

7 increase in protection compared to the then-current standard. Based on results of the exposure 

8 assessment, he also noted that exposures to O3 concentrations at and above a benchmark level of 

9 80 ppb would be essentially eliminated with a standard level of 75 ppb, and that exposures at and 

10 above a 70 ppb benchmark level would be substantially reduced or eliminated for the vast 

11 majority of people in at-risk groups. In addition, the Administrator concluded that the body of 

12 evidence did not support setting a lower standard level, specifically judging that the available 

13 evidence for effects following exposures to O3 concentrations of 60 ppb was “too limited to 

14 support a primary focus at this level” (75 FR 2938). With respect to the epidemiologic evidence, 

15 the Administrator stated that a standard set at a level lower than 75 ppb “would only result in 

16 significant further public health protection if, in fact, there is a continuum of health risks in areas 

17 with 8-hour average O3 concentrations that are well below the concentrations observed in the key 

18 controlled human exposure studies and if the reported associations observed in the 

19 epidemiological studies are, in fact, causally related to O3 at those lower levels” (73 FR 16483). 

20 In making his final decision about the level of the primary O3 standard, the Administrator 

21 noted that the level of 75 ppb was above the range recommended by CASAC (i.e., 70 to 60 ppb). 

22 He concluded that “CASAC’s recommendation appeared to be a mixture of scientific and policy 

23 considerations” (75 FR 2992). The Administrator reached a different policy judgment than the 

24 CASAC Panel, placing less weight than CASAC on the available controlled human exposure 

25 studies reporting effects following exposures to 60 ppb O3 and less weight on the results from 

26 exposure and risk assessments, particularly on estimates of exposures to O3 concentrations at or 

27 above 60 ppb (73 FR 16482-3). 

28 1.3.1.2 Approach for the Current Review 

29 To identify the range of options appropriate for the Administrator to consider in the 

30 current review, we apply an approach that builds upon the general approach used in the last 

31 review (and in the 2010 reconsideration proposal) and that reflects the broader body of scientific 

32 evidence, updated exposure/risk information, and advances in O3 air quality modeling now 

33 available. As summarized above, the Administrator’s decisions in the prior review were based on 

34 an integration of information on health effects associated with exposure to O3, judgments on the 
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1 adversity and public health significance of key health effects, and expert and policy judgments as 

2 to when the standard is requisite to protect public health with an adequate margin of safety. 

3 Staff’s preliminary conclusions on the primary O3 standard reflect our consideration of 

4 the available scientific evidence, exposure/risk information, and air quality modeling 

5 information, within the context of the overarching questions related to: (1) the adequacy of the 

6 current primary O3 standard to protect against effects associated with both short- and long-term 

7 exposures and (2) potential alternative standards, if any, that are appropriate to consider in this 

8 review. In addressing these broad questions, we organize the discussions in chapters 3 and 4 of 

9 this document around a series of more specific questions reflecting different aspects of each 

10 overarching question. When evaluating the health protection afforded by the current or potential 

11 alternative standards, we take into account the four basic elements of the NAAQS: the indicator, 

12 averaging time, form, and level. 

13 Figure 1-1 below provides an overview of our approach in this review. We believe that 

14 the general approach summarized in this section, and outlined in Figure 1-1, provides a 

15 comprehensive basis to help inform the judgments required of the Administrator in reaching 

16 decisions about the current and potential alternative primary O3 standards. In the subsections 

17 below, we describe our general approaches to considering the scientific evidence (evidence18 based considerations) and to considering the human exposure- and health risk information 

19 (exposure- and risk-based considerations). We also recognize considerations related to ambient 

20 O3 attributable to background sources. 
1-19

1

2 Figure 1-1. Overview of approach to reviewing the primary standard. 
1-20

1 1.3.1.2.1 Consideration of the Scientific Evidence 

2 Our approach in this review draws upon an integrative synthesis of the entire body of 

3 available scientific evidence for O3-related health effects, including the evidence newly available 

4 in the current review and the evidence from previous reviews, as presented in the ISA (U.S. 

5 EPA, 2013). Our approach to considering the scientific evidence is based fundamentally on using 

6 information from controlled human exposure and epidemiologic studies, supplemented by 

7 information from animal toxicology studies. Such evidence informs our consideration of the 

8 health endpoints and at-risk populations on which to focus the current review, and our 

9 consideration of the O3 concentrations at which various health effects can occur. 

10 Since the 2008 review of the O3 NAAQS, the Agency has developed formal frameworks 

11 for characterizing the strength of the scientific evidence with regard to health effects associated 

12 with exposures to O3 in ambient air and factors that may increase risk in some populations or 

13 lifestages (U.S. EPA, 2013, Preamble; Chapter 8). These frameworks provide the basis for 

14 robust, consistent, and transparent processes for evaluating the scientific evidence, including 

15 uncertainties in the evidence, and for drawing weight-of-evidence conclusions on air pollution16 related health effects and at-risk populations. 

17 With regard to characterization of health effects, the ISA uses a five-level hierarchy to 

18 classify the overall weight-of-evidence into one of the following categories: causal relationship, 

19 likely to be a causal relationship, suggestive of a causal relationship, inadequate to infer a causal 

20 relationship, and not likely to be a causal relationship (U.S. EPA, 2013, Preamble Table II). In 

21 this PA, we place the greatest weight on the evidence for health effects that have been judged in 

22 the ISA to be caused by, or likely to be caused by, O3 exposures. Our consideration of the 

23 available evidence for such effects is presented below in Chapter 3 (consideration of the 

24 adequacy of the current standard) and in Chapter 4 (consideration of potential alternative 

25 standards).

26 As discussed below, we further consider the evidence base assessed in the ISA with 

27 regard to the types and levels of exposure at which health effects are indicated. This further 

28 consideration of the evidence, which directly informs EPA’s conclusions regarding the adequacy 

29 of current or potential alternative standards in providing requisite public health protection, differs 

30 from consideration of the evidence in the ISA with regard to overarching determinations of 

31 causality. Therefore, studies that inform determinations of causality may or may not be 

32 concluded to be informative with regard to the adequacy of the current or potential alternative 

33 standards. 

34 As with health endpoints, the ISA’s characterization of the weight-of-evidence for 

35 potential at-risk populations is based on the evaluation and synthesis of evidence from across 

36 scientific disciplines. The ISA characterizes the evidence for a number of “factors” that have the 
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1 potential to place populations at increased risk for O3-related effects. The categories considered 

2 in evaluating the evidence for these potential at-risk factors are “adequate evidence,” “suggestive 

3 evidence,” “inadequate evidence,” and “evidence of no effect.” These categories are discussed 

4 in more detail in the ISA (U.S. EPA, 2013, chapter 8, Table 8-1). In this draft PA, we focus our 

5 consideration of potential at-risk populations on those factors for which the ISA judges there is 

6 “adequate” evidence (U.S. EPA, 2013, Table 8-5). At-risk populations are discussed in more 

7 detail in section 3.2.1, below. 

8 Using the available scientific evidence to inform conclusions on the adequacy of the 

9 current primary O3 standard, and on potential alternative standards appropriate for consideration, 

10 is complicated by the recognition that a population-level threshold in exposure or ambient O3

11 concentrations has not been identified, below which it can be concluded with confidence that O3-

12 attributable effects do not occur in exposed populations (U.S. EPA, 2013, section 2.5.4.4). In the 

13 absence of a discernible threshold, our general approach to considering the available O3 health 

14 evidence involves characterizing our confidence in the extent to which O3-attributable effects 

15 occur, and the extent to which such effects are adverse, over the ranges of O3 exposure 

16 concentrations evaluated in controlled human exposure studies and over the distributions of 

17 ambient O3 concentrations in locations where epidemiologic studies have been conducted. As 

18 noted above, we recognize that the available health effects evidence reflects a continuum from 

19 relatively high O3 concentrations, at which scientists generally agree that adverse health effects 

20 are likely to occur, through lower concentrations, at which the likelihood and magnitude of a 

21 response become increasingly uncertain. Aspects of our approach particular to evidence from 

22 controlled human exposure and epidemiologic studies, respectively, are discussed below. 

23 Controlled Human Exposure Studies 

24 Controlled human exposure studies provide direct evidence of relationships between 

25 pollutant exposures and human health effects (U.S. EPA, 2013, p.lx). Such studies are 

26 particularly useful in defining the specific conditions under which pollutant exposures can result 

27 in health impacts, including the exposure concentrations, durations, and ventilation rates under 

28 which effects can occur. As discussed in the ISA, controlled human exposure studies provide 

29 clear and compelling evidence for an array of human health effects that are directly attributable 

30 to acute exposures to O3 per se (i.e., as opposed to O3 and other photochemical oxidants, for 

31 which O3 is an indicator, or other co-occurring pollutants) (U.S. EPA, 2013, Chapter 6). 

32 Together with animal toxicological studies, which can provide information about more serious 

33 health outcomes as well as the effects of long-term exposures and mode of action, controlled 

34 human exposure studies also help to provide biological plausibility for health effects observed in 

35 epidemiologic studies. 
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1 In this draft PA, we consider the evidence from controlled human exposure studies in two 

2 ways. First, we consider the extent to which controlled human exposure studies provide evidence 

3 for health effects following exposures to different O3 concentrations, down to the lowest4 observed-effects levels in those studies. Second, we use such studies to inform our evaluation of 

5 the extent to which we have confidence in health effect associations reported in epidemiologic 

6 studies down through lower ambient O3 concentrations, where the likelihood and magnitude of 

7 O3-attributable effects become increasingly uncertain. 

8 We consider the range of O3 exposure concentrations evaluated in controlled human 

9 exposure studies, including concentrations near or below the level of the current standard. We 

10 consider both group mean responses, which provide insight into the extent to which observed 

11 changes are due to O3 exposures rather than to chance alone, and inter-individual variability in 

12 responses, which provides insight into the fraction of the population that might be affected by 

13 such O3 exposures (U.S. EPA, 2013, section 6.2.1.1). When considering the relative weight to 

14 place on various controlled human exposure studies, we consider the exposure conditions 

15 evaluated (e.g., exercising versus resting, exposure duration); the nature, magnitude, and likely 

16 adversity of effects over the range of reported O3 exposure concentrations; the statistical 

17 precision of reported effects; and the consistency of results across studies for a given health 

18 endpoint and exposure concentration. In addition, because controlled human exposure studies 

19 typically involve healthy individuals and do not evaluate the most sensitive individuals in the 

20 population (U.S. EPA, 2013, Preamble p. lx), when considering the implications of these studies 

21 for our evaluation of the current and potential alternative standards, we also consider the extent 

22 to which reported effects are likely to reflect the magnitude and/or severity of effects in at-risk 

23 groups. 

24 Epidemiologic Studies 

25 We also consider epidemiologic studies of short- and long-term O3 concentrations in 

26 ambient air. Epidemiologic studies provide information on associations between variability in 

27 ambient O3 concentrations and variability in various health outcomes, including lung function 

28 decrements, respiratory symptoms, school absences, hospital admissions, emergency department 

29 visits, and premature mortality (U.S. EPA, 2013, Chapters 6 and 7). Epidemiologic studies can 

30 inform our understanding of the effects in the study population (which may include at-risk 

31 groups) of real-world exposures to the range of O3 concentrations in ambient air. 

32 Available studies have generally not indicated a discernible population threshold, below 

33 which O3 is no longer associated with health effects (U.S. EPA, section 2.5.4.4). However, the 

34 currently available epidemiologic evidence indicates decreased confidence in reported 

35 concentration-response relationships for O3 concentrations at the lower ends of ambient 

36 distributions (U.S. EPA, section 2.5.4.4). Therefore, our general approach to considering the 
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1 results of epidemiologic studies within the context of the current and potential alternative 

2 standards focuses on characterizing the range of ambient O3 concentrations over which we have 

3 the most confidence in O3-associated health effects, and the concentrations below which our 

4 confidence in such health effect associations becomes appreciably lower. In doing so, we 

5 consider the statistical precision of O3 health effect associations reported in study locations with 

6 various ambient O3 concentrations; confidence intervals around concentration-response functions 

7 reported over distributions of ambient O3 (where available); and the extent to which the 

8 biological plausibility of associations at various ambient O3 concentrations is supported by 

9 evidence from controlled human exposure and/or animal toxicological studies. 

10 We consider both multi-city and single-city studies assessed in the ISA, each of which 

11 have strengths and limitations. Multi-city studies evaluate large populations and provide greater 

12 statistical power than single-city studies. Multi-city studies also reflect O3-associated health 

13 impacts across a range of diverse locations, providing spatial coverage for different regions 

14 across the country and reflecting differences in exposure-related factors that could impact O3

15 risks. In addition, compared to single-city studies, multi-city studies are not prone to publication 

16 bias and they afford the possibility of generalizing to the broader national population (U.S. EPA, 

17 2004, p. 8-30). In contrast, while single-city studies are more limited than multicity studies in 

18 terms of statistical power and geographic coverage, conclusions regarding the extent to which air 

19 quality met the current or potential alternative standards in the cities for which associations have 

20 been reported can be made with greater certainty for single-city studies (compared to multicity 

21 studies reporting only multicity effect estimates) because the associations are reported for cityspecific analyses (U.S. EPA, 2011d, section 2.3.4.1).19 22 In some cases, single-city studies can 

23 also provide evidence for locations or population-specific characteristics not reflected in 

24 multicity studies (U.S. EPA, 2013, section 6.2.7.1). Therefore, when considering available 

25 epidemiologic studies we evaluate both multi-city and single-city studies, recognizing the 

26 strengths and limitations of each. 

27 In placing emphasis on specific epidemiologic studies, we focus on studies conducted in 

28 the U.S. and Canada. Such studies reflect air quality and exposure patterns that are likely more 

29 typical of the U.S. population than the air quality and exposure patterns reflected in studies 

conducted outside the U.S. and Canada.20 30 We also focus on studies reporting associations with 

31 effects judged in the ISA to be robust to confounding by other factors, including co-occurring air 

32 pollutants. 

 

19 Though in some cases multicity studies present single-city effect estimates in addition to multi-city estimates. 

20 All studies, including other international studies inform the causal determinations in the ISA. 
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1 1.3.1.2.2 Consideration of Exposure and Risk Estimates 

2 To put judgments about O3-related health effects into a broader public health context, we 

3 consider exposure and risk estimates from the second draft HREA, which develops and applies 

4 models to estimate human exposures to O3 and O3-related health risks in urban case study areas 

5 across the United States (U.S. EPA, 2014). The second draft HREA estimates exposures of 

6 concern, based on interpreting quantitative exposure estimates within the context of controlled 

7 human exposure study results; lung function risks, based on applying exposure-response 

8 relationships from controlled human exposure studies to quantitative estimates of exposures; and 

9 epidemiologic-based risk estimates, based on applying concentration-response relationships 

10 drawn from epidemiologic studies to adjusted air quality. Each of these types of assessments is 

11 discussed briefly below. 

12 As in the 2008 review, the second draft HREA estimates exposures at or above 

13 benchmark concentrations of 60, 70, and 80 ppb, reflecting exposure concentrations of concern 

based on the available health evidence.21 14 Estimates of exposures at or above discrete benchmark 

15 concentrations provide perspective on the public health risks of O3-related health effects that 

16 have been demonstrated in controlled human exposure and toxicological studies but that, because 

17 of a lack of exposure-response information from those studies, cannot be assessed using a 

18 quantitative risk assessment. Though this analysis is conducted using discrete benchmark 

19 concentrations, health-relevant exposures are more appropriately viewed as a continuum with 

20 greater confidence and less uncertainty about the existence of health effects at higher O3

21 exposure concentrations and less confidence and greater uncertainty at lower exposure 

22 concentrations. We recognize that there is no sharp breakpoint within the exposure-response 

23 relationship for exposure concentrations at and above 80 ppb down to 60 ppb. 

24 The second draft HREA also generates quantitative estimates of O3 health risks for air 

25 quality adjusted from recent conditions to those just meeting the current and potential alternative 

26 standards. As noted above, one approach to estimating O3 health risks is to combine modeled 

27 exposure estimates with exposure-response relationships derived from controlled human 

28 exposure studies of O3-induced health effects. The second draft HREA uses this approach to 

29 estimate the occurrence of O3-induced lung function decrements in simulated at-risk populations. 

30 The available exposure-response information does not support this approach for other endpoints 

31 evaluated in controlled human exposure studies (U.S. EPA, 2014a, section 2.3). 

32 Another approach to estimating O3-associated health risks is to apply concentration33 response relationships derived from short- and/or long-term epidemiologic studies to air quality 

 

21 For example, see 75 FR 2945-2946 (January 19, 2010) and 73 FR 16441-16442 (March 27, 2008) discussing 

“exposures of concern”. 
1-25

1 adjusted to just meet current and potential alternative standards. The concentration-response 

2 relationships drawn from epidemiologic studies are based on population exposure surrogates, 

3 such as 8-hour concentrations averaged across monitors and over more than one day 

4 (incorporation of lag) (U.S. EPA, 2013, Chapter 6). The second draft HREA presents 

5 epidemiologic-based risk estimates for O3-associated mortality, hospital admissions, emergency 

6 department visits, and respiratory symptoms (U.S. EPA, 2014a, section 2.3). These estimates are 

7 derived from the full distribution of ambient O3 concentrations estimated for the study 

locations.22 8 In addition, the second draft HREA estimates mortality risks attributable to various

9 portions of those distributions (U.S. EPA, 2014a). In this second draft PA we consider risk 

10 estimates based on the full distributions of ambient O3 concentrations, and estimates of the risk 

11 associated with various portions of those ambient distributions. In doing so, we take note of the 

12 ISA conclusions regarding confidence in linear concentration-response relationships over 

13 distributions of ambient concentrations, and of the extent to which health effect associations at 

14 various ambient O3 concentrations are supported by the evidence from experimental studies for 

15 effects following specific O3 exposures. 

16 1.3.1.2.3 Considerations Regarding Ambient O3 Concentration Estimates 

17 Attributable to Background Sources 

18 As noted above, our approach in this review utilizes recent advances in modeling 

19 techniques to estimate the contributions of U.S. anthropogenic, international anthropogenic, and 

20 natural sources to ambient O3 (discussed in detail in Chapter 2 of this draft PA). Such model 

21 estimates can provide insights into the extent to which different types of background emissions 

22 sources contribute to total ambient O3 concentrations. Consideration of this issue in the current 

23 review is informed by the approaches taken in previous reviews, as well as by court decisions. 

24 In 1979, the EPA set a 1-hour O3 standard with a level of 0.12 ppm. Following the final 

25 decision in that review, the City of Houston argued that the standard was arbitrary and capricious 

26 because natural O3 concentrations and other physical phenomena in the Houston area made the 

27 standard unattainable in that area. The D.C. Circuit rejected this argument, stating that 

28 attainability and technological feasibility are not relevant considerations in the promulgation of 

29 the NAAQS. The Court also noted that the EPA need not tailor the NAAQS to fit each region or 

30 locale, pointing out that Congress was aware of the difficulty in meeting standards in some 

 

22 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).
1-26

1 locations and had addressed this difficulty through various compliance related provisions in the 

2 Act. See API v. Costle, 665 F.2d 1176, 1184-6 (D.C. Cir. 1981). 

3 More recently, in the 1997 review of the O3 NAAQS, the Administrator set an 8-hour 

4 standard with a level of 0.08 ppm (84 ppb). In reaching this decision, the EPA identified several 

5 reasons supporting its decision to reject a more stringent standard of 0.07 ppm. Most 

6 importantly, the EPA pointed out the scientific uncertainty at lower concentrations and placed 

7 significant weight on the fact that no CASAC panel member supported a standard level set lower 

8 than 0.08 ppm (62 FR 38868). In addition to noting the uncertainties in the health evidence for 

9 exposure concentrations below 0.08 ppm and the advice of CASAC, the EPA noted that a 

10 standard set at a level of 0.07 ppm would be closer to peak background concentrations that 

11 infrequently occur in some areas due to nonanthropogenic sources of O3 precursors (62 FR 

12 38856, 38868; July 18, 1997). 

13 In subsequent litigation, the D.C. Circuit upheld the EPA’s decision as the product of 

14 reasoned decision-making. The Court made clear that the most important support for the EPA’s 

15 decision was the health evidence and the concerns it raised about setting a standard level below 

16 0.08 ppm. The Court also pointed to the significant weight that the EPA properly placed on the 

17 advice it received from CASAC. Finally (as noted in section 1.2.2 above), the Court noted that 

18 the EPA could also consider relative proximity to peak natural background O3 when evaluating 

19 alternative standards. See ATA III, 283 F.3d at 379 (D.C. Cir. 2002). 

20 These cases provide a framework for considering the contributions of U.S. 

21 anthropogenic, international anthropogenic, and natural sources, within the context of 

22 considering the health evidence and CASAC advice, when evaluating various potential 

23 alternative standards. 

24 1.3.2 Approach for the Secondary Standard 

25 Staff’s approach in this review of the current secondary standard takes into consideration 

26 aspects of the approaches used in past O3 NAAQS reviews. The past and current approaches, 

27 generally described below, are both based fundamentally on using EPA’s assessment of the 

28 current scientific evidence and associated quantitative analyses to inform the Administrator’s 

29 judgment regarding a secondary standard for O3 that is requisite (i.e., neither more nor less 

30 stringent than necessary) to protect public welfare. 

31 In reaching conclusions on options for the Administrator’s consideration, we note that the 

32 final decision to retain or revise the current secondary O3 standard is a public welfare policy 

33 judgment to be made by the Administrator. This final decision will draw upon the available 

34 scientific evidence for O3-attributable welfare effects and on analyses of vegetation and 

35 ecosystem exposures and public welfare risks based on impacts to vegetation, ecosystems and 
1-27

1 their associated services, including judgments about the appropriate weight to place on the range 

2 of uncertainties inherent in the evidence and analyses. In determining the requisite level of 

3 protection for crops and trees, the Administrator will need to weigh the importance of the 

4 predicted risks of these effects in the overall context of public welfare protection, along with a 

5 determination as to the appropriate weight to place on the associated uncertainties and limitations 

6 of this information. Our general approach to informing these judgments, discussed more fully 

7 below, recognizes that the available welfare effects evidence reflects a continuum from relatively 

8 high O3 concentrations at which scientists generally agree that welfare effects are likely to occur, 

9 through lower concentrations at which the likelihood and magnitude of a response become 

10 increasingly uncertain. Therefore, in developing conclusions in this second draft PA, we are 

11 mindful that the Administrator’s ultimate judgments on the secondary standard will most 

12 appropriately reflect an interpretation of the available scientific evidence and exposure/risk 

13 information that neither overstates nor understates the strengths and limitations of that evidence 

14 and information. 

15 Section 1.3.2.1 below provides an overview of the general approach taken in the last 

16 review of the secondary standard for O3 (i.e., the 2008 review), and a summary of the rationale 

17 for the decision on the standard in that review (73 FR 16436). Section 1.3.2.2 presents our 

18 approach in the current review, including our approach to considering the vegetation effects 

19 evidence and exposure/risk information, and considerations regarding ambient O3 concentrations 

20 attributable to background sources. 

21 1.3.2.1 Approach Used in the Last Review 

22 In the 2008 review of the secondary NAAQS for O3, the Administrator relied upon 

23 consideration of the available scientific evidence and exposure/risk information, information 

24 regarding biologically-relevant exposure indices, air quality information regarding the degree of 

25 overlap between different exposure index forms, the advice and recommendations of CASAC, 

26 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 ppm23 to 0.075 ppm (75 ppb24 27 ). 

28 In reaching a decision to revise the 1997 8-hour secondary standard, the Administrator 

29 found, after carefully considering the public comments, that the fundamental scientific 

30 conclusions on the effects of O3 on vegetation and sensitive ecosystems reached in the 2006 

31 Criteria Document and 2007 Staff Paper, as discussed in section IV.A of the final rule remained 

 

23 Due to rounding convention, the 1997 standard level of 0.08 ppm corresponded to 0.084 ppm (84 ppb). 

24 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, n order to avoid the confusion that could 

result from switching units when discussing the evidence in relation to the standard level. 
1-28

1 valid (73 FR 16496). He further recognized that several additional lines of evidence had 

2 progressed sufficiently since the 1997 review to provide a more complete and coherent picture of 

3 the scope of O3-related vegetation risks (i.e., visible foliar injury, tree biomass loss, crop yield 

4 loss, and others), especially those faced by sensitive seedling, sapling and mature growth stage 

5 tree species growing in field settings, and their associated forested ecosystems. This new 

6 research reflected an increased emphasis on field-based exposure methods (e.g., free-air, ambient 

7 gradient and biomonitoring surveys) (73 FR 16490) in addition to the more traditional controlled 

8 open-top chamber (OTC) studies (73 FR 16485), and began to address one of the key data gaps 

9 cited by the Administrator in the 1997 review (73 FR 16486). Specifically, by providing 

10 additional evidence that O3-induced crop yield loss and tree seedling biomass loss effects 

11 observed in chambers also occurs in the field, this new research qualitatively increased support 

12 for, and confidence in, the continued use of OTC-derived crop and tree seedling concentration13 response (C-R) functions developed in the National Crop Loss Assessment Network (NCLAN) 

14 and National Health and Environmental Effects Research Laboratory – Western Ecology 

15 Division (NHEERL-WED) studies, respectively, to predict O3-induced impacts on crops and tree 

16 seedlings in the field (72 FR 37886). All of these areas were considered together, along with 

17 associated uncertainties, in an integrated weight-of-evidence approach (73 FR 16490). 

18 Beyond the available vegetation effects evidence, the Administrator also considered 

19 estimates of O3 exposures and risks when air quality was adjusted to simulate just meeting the 

20 existing and potential alternative standards. On the basis of these assessments, the Administrator 

21 concluded that O3 exposures that would be expected to remain after meeting the existing 

22 standard would be sufficient to cause visible foliar injury and seedling and mature tree biomass 

23 loss in O3-sensitive vegetation (73 FR 16496) and would still allow O3-related yield loss to occur 

24 in some commodity crop species and fruit and vegetable species grown in the U.S. (73 FR 

25 16489). Other O3-induced effects described in the literature, including an impaired ability of 

26 many sensitive species and genotypes within species to adapt to or withstand other 

27 environmental stresses, such as freezing temperatures, pest infestations and/or disease, and to 

28 compete for available resources, would also be anticipated to occur. In the long run, the result of 

29 these impairments (e.g., loss in vigor) could lead to premature plant death in O3 sensitive species. 

30 Though effects on other ecosystem components had only been examined in isolated cases, the 

31 Administrator noted effects such as those described above could have significant implications for 

32 plant community and associated species biodiversity and the structure and function of whole 

33 ecosystems (73 FR 16496). 

34 Although the Administrator concluded that the then-current standard was not sufficient to 

35 protect against the known and anticipated effects described above, he also recognized that the 

36 secondary standard is not meant to protect against all known observed or anticipated O3-related 
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1 effects, but only those that can reasonably be judged to be adverse to the public welfare. The 

2 Administrator recognized that the degree to which such effects should be considered to be 

3 adverse depended on the intended use of the vegetation and its significance to the public welfare 

4 (73 FR 16496). In this regard, he took note of a number of actions taken by Congress to establish 

5 public lands that are set aside for specific uses that are intended to provide benefits to the public 

6 welfare, including lands that are to be protected so as to conserve the scenic value and the natural 

7 vegetation and wildlife within such areas, and to leave them unimpaired for the enjoyment of 

8 future generations. Based on these considerations, and taking into consideration the advice and 

9 recommendations of CASAC, the Administrator concluded that the protection afforded by the 

10 existing standard was not sufficient, and that the standard needed to be revised to provide 

11 additional protection from known and anticipated adverse effects on sensitive natural vegetation 

12 and ecosystems (73 FR 16497). 

13 Given his judgment on the need to revise, the Administrator then considered what 

14 revisions to the standard were requisite to protect public welfare. Regarding the form of the 

15 standard, the Administrator took note that at the conclusion of the 1997 review, the biological 

basis for a cumulative, seasonal form was not in dispute25 16 and that the 2006 Criteria Document 

17 also concluded that O3 exposure indices that cumulate differentially-weighted hourly 

18 concentrations are the best candidates for relating exposure to plant growth responses (EPA, 

19 2006) (61 FR 65716; 73 FR 16486). The CASAC, in its letter to the Administrator following its 

20 review of the second draft Staff Paper, stated that “there is a clear need for a secondary standard 

21 which is distinctly different from the primary standard in averaging time, level and form” and 

22 that “the CASAC unanimously agrees that it is not appropriate to try to protect vegetation from 

23 the substantial, known or anticipated, direct and/or indirect, adverse effects of ambient ozone by 

24 continuing to promulgate identical primary and secondary standards for ozone” (Henderson, 

25 October 24, 2006, pp. 5-7). Although many possible cumulative, seasonal concentration26 weighted exposure metrics exist, the Staff Paper and the CASAC Panel concluded that the 

W12626 27 form is the most biologically-relevant cumulative, seasonal form appropriate to consider 

in the context of the secondary standard review (73 FR 16486-87).27 28 

 

25 In the 1997 review, a different cumulative metric (SUM06) was proposed. Metric selection in both 1997 and 2008 

was based on both science and policy considerations. 

26 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 sigmoidally 

weighted hourly O3 concentrations over a specified period, where the daily sigmoidal weighting function is defined 

as: 1 – exp[-(W126/η)β]

27 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 
1-30

1 Although agreeing with the Criteria Document, Staff Paper and CASAC conclusions that 

2 a cumulative exposure index that differentially weights O3 concentrations could represent a 

3 reasonable policy choice for a seasonal secondary standard to protect against the effects of O3 on 

4 vegetation and that the most appropriate cumulative, concentration-weighted form to consider 

5 was the sigmoidally weighted W126 form (73 FR 16498), the Administrator also took note of the 

6 1997 decision to make the revised secondary standard identical to a revised primary standard 

7 after similar considerations (73 FR 16498). In considering the rationale for the 1997 decision, the 

8 Administrator observed that it was based in part on an analysis that compared the degree of 

9 overlap in county-level air quality measured in terms of alternative standard forms (62 FR 

10 38876). Recognizing that significant uncertainty remained in 1997 regarding conclusions drawn 

11 from such analyses, the Administrator also considered the results of a similar analysis of recent 

12 monitoring data undertaken in the 2007 Staff Paper to assess the degree of overlap expected 

between the existing standard (4th 13 high, daily maximum 8-hour concentration averaged over 

14 three years) and potential alternative standards based on W126 cumulative seasonal forms. 

15 The Administrator noted that this analysis showed significant overlap between the 8-hour 

16 secondary standard and selected levels of W126 standard forms, with the degree of overlap 

17 between these potential alternative standards depending greatly on the W126 level selected and 

18 the distribution of hourly O3 concentrations within the annual and/or 3-year average period. 

19 From this analysis, the Administrator recognized that a secondary standard set identical to a 

20 revised primary standard would provide a significant degree of additional protection for 

21 vegetation as compared to that provided by the existing secondary standard. In further 

22 considering the significant uncertainties in the available body of evidence and in the exposure 

23 and risk analyses, and the difficulty in determining at what point various types of vegetation 

24 effects become adverse for sensitive vegetation and ecosystems, the Administrator focused his 

25 consideration on a level for an alternative W126 standard (with an annual form) at the upper end 

26 of the proposed range (i.e., 21 ppm-hours). The Staff Paper analysis showed that at a W126 level 

27 of 21 ppm-hours, there would be essentially no counties with air quality expected both to exceed 

28 such an alternative W126 standard and to meet the revised 8-hour primary standard—that is, 

29 based on this analysis of counties with ambient O3 monitors, a W126-based level of 21 ppm30 hours would be unlikely to provide additional protection in any areas beyond that likely to be 

31 provided by the revised 2008 primary standard (73 FR 16499/500). 

 

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). 
1-31

1 The Administrator also considered the Staff Paper finding that the degree of overlap 

2 between counties (with areas of concern for vegetation) expected to meet an 8-hour level for the 

3 form of the existing standard and potential alternative levels of a W126-based standard was 

4 inconsistent across years analyzed. This variation depended greatly on levels selected for a 

W126-based standard and a 3-year average 4th 5 high daily maximum 8-hour standard, 

6 respectively, and the distribution of hourly O3 concentrations within the annual and/or 3-year 

7 average period. From this, the Staff Paper recognized the need for caution in evaluating the 

8 likely vegetation impacts associated with a given level of air quality expressed in terms of the 

9 existing 8-hour average standard in the absence of parallel W126 information. In considering 

10 these findings, the Administrator “recognize[d] that the general lack of rural monitoring data 

11 made uncertain the degree to which the revised 8-hour standard or an alternative W126 standard 

12 would be protective, and that there was the potential for not providing the appropriate degree of 

13 protection for vegetation in areas with air quality distributions that resulted in a high cumulative, 

14 seasonal exposure but did not result in high 8-hour average exposures” (73 FR 16500). With 

15 regard to the 8-hour standard, he also noted that “[w]hile this potential for under-protection was 

16 clear, the number and size of areas [then] at issue and the degree of risk [was] hard to determine. 

17 However, such a standard would also tend to avoid the potential for providing more protection 

18 than is necessary, a risk that would have arisen from moving to a new form for the secondary 

19 standard despite the significant uncertainty in determining the degree of risk for any exposure 

20 level and the appropriate level of protection, as well as uncertainty in predicting exposure and 

21 risk patterns” (73 FR 16500). 

22 Thus, although the Administrator agreed with the views and recommendations of 

23 CASAC that a cumulative, seasonal standard was the most biologically relevant way to relate 

24 exposure to plant growth response, he also recognized that there remained significant 

25 uncertainties in determining or quantifying the degree of risk attributable to varying levels of O3

26 exposure, the degree of protection that any specific cumulative, seasonal standard would 

27 produce, and the associated potential for error in determining the secondary standard that would 

28 provide a requisite degree of protection—i.e., sufficient but not more than what is necessary. 

29 Given these significant uncertainties, the Administrator concluded that establishing a new 

30 secondary standard with a cumulative, seasonal form, at that time, would have resulted in 

31 uncertain benefits beyond those afforded by the revised primary standard, and therefore, might 

32 have been more than necessary to provide the requisite degree of protection (73 FR 16500). 

33 Based on his consideration of these issues (73 FR 16497), the Administrator judged that the 

34 appropriate balance to be drawn was to set a secondary standard identical in every way to the 

35 revised 8-hour primary standard of 0.075 ppm. The Administrator believed that such a standard 

36 would be sufficient to protect public welfare from known or anticipated adverse effects, and did 
1-32

1 not believe that an alternative cumulative, seasonal standard was needed to provide this degree of 

2 protection (73 FR 16500). 

3 As noted above, on July 23, 2013 the D.C. Circuit found this approach to be contrary to 

4 law because EPA had failed to identify a level of air quality requisite to protect public welfare 

5 and, therefore, EPA’s comparison between the primary and secondary standards for determining 

6 if requisite protection for public welfare was afforded by the primary standard was inherently 

7 arbitrary. The court remanded the secondary standard to EPA for further consideration. 723 F. 3d 

8 at 270-74. 

9 1.3.2.2 Approach for the Current Review 

10 To identify the range of options appropriate for the Administrator to consider in the 

11 current review, we apply an approach that builds upon the general approach used in the 2008 

12 review(and in the 2010 reconsideration proposal), and that reflects the broader body of scientific 

13 evidence, updated exposure/risk information, and advances in O3 air quality modeling now 

14 available. As summarized above, the Administrator’s decisions in the prior review were based on 

15 an integration of information on welfare effects associated with exposure to O3, judgments on the 

16 adversity and public welfare significance of key effects, and, expert and policy judgments as to 

17 when the standard is requisite to protect public welfare. These considerations were informed by 

18 air quality and related analyses, quantitative exposure and risk assessments, and qualitative 

19 assessment of impacts that could not be quantified. In performing the evaluation in this 

20 document, we are additionally mindful of the recent remand of the secondary standard by the 

21 D.C. Circuit. 

22 Our approach in this review of the secondary O3 standard reflects our consideration of the 

23 available scientific evidence, information on biologically-relevant exposure indices, 

24 exposure/risk information, and air quality modeling information, within the context of 

25 overarching questions related to: (1) the adequacy of the current secondary O3 standard to protect 

26 against effects associated with cumulative, seasonal exposures and (2) potential alternative 

27 standards, if any, that are appropriate to consider in this review. In addressing these broad 

28 questions, we have organized the discussions in chapters 5 and 6 of this document around a 

29 series of more specific questions reflecting different aspects of each overarching question. When 

30 evaluating the welfare protection afforded by the current or potential alternative standards, we 

31 take into account the four basic elements of the NAAQS: the indicator, averaging time, form, 

32 and level. 

33 Figure 1-2 below provides an overview of our approach in this review. We believe that 

34 the general approach summarized in this section, and outlined in Figure 1-2, provides a 

35 comprehensive basis to help inform the judgments required of the Administrator in reaching 
1-33

1 decisions about the current and potential alternative secondary O3 standards. In the subsections 

2 below, we summarize our general approaches to considering the scientific evidence (evidence3 based considerations) and to considering the exposure and risk information (exposure- and risk4 based considerations). We also recognize considerations related to ambient O3 attributable to 

5 background sources. 
1-34

1

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

1 1.3.2.2.1 Consideration of the Scientific Evidence 

2 Our approach in this review draws upon an integrative synthesis of the entire body of 

3 available scientific evidence for O3-related welfare effects, including the evidence newly 

4 available in the current review and the evidence from previous review, as presented in the ISA 

5 (U.S. EPA, 2013). Our approach to considering the scientific evidence for effects on vegetation 

6 is based fundamentally on using information from controlled chamber studies and field-based 

7 studies. Such evidence informs our consideration of welfare endpoints and at-risk species and 

8 ecosystems on which to focus the current review, and our consideration of the ambient O3

9 conditions under which various welfare effects can occur. 

10 As in each NAAQS review, we consider the entire body of evidence for the subject 

11 criteria pollutant. With regard to identification of the welfare effects that could be caused by a 

12 pollutant, we look to controlled exposure studies using chamber or free air methodologies and 

13 field-based observational, survey and gradient studies. Evaluating all of the evidence together, 

14 the ISA makes a determination with regard to the strength of the evidence for a causal 

15 relationship between the air pollutant and specific welfare effects. These determinations inform 

16 our identification of welfare effects for which the NAAQS may provide protection. 

17 Since the 2008 review of the O3 NAAQS, the Agency has developed a formal framework 

18 for characterizing the strength of the scientific evidence with regard to a causal relationship 

19 between ambient O3 and welfare effects (U.S. EPA, 2013, Preamble; Chapter 9). This framework 

20 provides the basis for a robust, consistent, and transparent process for evaluating the scientific 

21 evidence, including uncertainties in the evidence, and for drawing weight-of-evidence 

22 conclusions regarding air pollution-related welfare effects. In so doing, the ISA uses a five-level 

23 hierarchy, classifying the overall weight of evidence into one of the following categories: causal 

24 relationship, likely to be a causal relationship, suggestive of a causal relationship, inadequate to 

25 infer a causal relationship, and not likely to be a causal relationship (U.S. EPA, 2013, Preamble 

26 Table II). In our approach here, we place the greatest weight on the evidence for welfare effects 

27 that have been judged in the ISA to be caused by, or likely caused by, O3 exposures. Our 

28 consideration of the available evidence for such effects is presented below in Chapter 5 

29 (consideration of the adequacy of the current standard) and in Chapter 6 (consideration of 

30 potential alternative standards).

31 We further consider the evidence base, as assessed in the ISA, with regard to the types 

32 and levels of exposure at which welfare effects are indicated. This further consideration of the 

33 evidence base, which directly informs EPA’s conclusions regarding the adequacy of current or 

34 potential alternative standards in providing requisite public welfare protection, differs from 

35 consideration of the evidenc in the ISA with regard to overarching determinations of causality. 
1-36

1 Studies that have informed determinations of causality may or may not be concluded to be 

2 informative with regard to the adequacy of the current or potential alternative standards. 

3 Our approach in this review, as in past reviews, included recognition that the available 

4 evidence has not provided identification of a threshold in exposure or ambient O3 concentrations 

5 below which it can be concluded with confidence that O3-attributable vegetation effects do not 

6 occur across the broad range of O3-sensitive plant species growing within the U.S. This is due in 

7 part to the fact that research shows that there is variability in sensitivity between and within 

8 species and that numerous factors, i.e. chemical, physical, biological, and genetic, can influence 

9 the direction and magnitude of the studied effect (U.S. EPA, 2013, section 9.4.8). In the absence 

10 of a discernible threshold, our general approach to considering the available O3 welfare evidence 

11 involves characterizing our confidence in conclusions regarding O3-attributable vegetation 

12 effects over the ranges of cumulative seasonal O3 exposure values evaluated in chamber studies 

13 and in field studies in areas where O3-sensitive vegetation are known to occur, as well as 

14 characterizing the extent to which these effects can be considered adverse. In addition, because 

15 O3 can indirectly affect other ecosystem components (such as soils, water, and wildlife, and their 

16 associated goods and services, through its effects on vegetation) our approach also considers 

17 those indirect effects for which the ISA concludes, based on multiple lines of evidence, including 

18 mechanistic and physiological processes, to have a causal or likely to be a causal relationship. 

19 With respect to ecosystem services for which we may have only limited or qualitative 

20 information regarding an association with O3 exposures, our approach is to consider their policy21 relevance in the context of section 109(b)(2) of the CAA which specifies that secondary 

22 standards provide requisite protection of “public welfare from any ... known or anticipated 

23 adverse effects associated with the presence of [the] pollutant in the ambient air”. As noted 

24 above, our approach recognizes that the effects evidence reflects a general continuum from 

25 higher O3 concentrations, at which scientists generally agree that adverse vegetation and 

26 ecosystem effects are likely to occur, through lower concentrations, at which the likelihood and 

27 magnitude of a response becomes increasingly more uncertain. 

28 In this review, the evidence base includes quantitative information across a broad array of 

29 vegetation effects (e.g., growth impairment during seedlings, saplings and mature tree growth 

30 stages, visible foliar injury, and yield loss in annual crops) and across a diverse set of exposure 

31 methods from laboratory and field studies. These methods include the more traditional OTC

32 studies, as well as field-based exposure studies. While we consider the full breadth of 

33 information available, we place greater weight on U.S. studies due to the often species-, site-, 

34 and climate-specific nature of O3-related vegetation responses. We especially weight those 

35 studies that include O3 exposures that fall within the range of those likely to occur in the ambient 

36 air. Further, our approach in the context of the quantitative exposure and risk assessments 
1-37

1 (discussed below), places greatest emphasis on studies that have evaluated plant response over 

2 multiple exposure levels and developed exposure-response relationships that allow the prediction 

3 (estimation) of plant responses over the range of potential alternative standards being assessed. 

4 In considering the evidence, we recognize differences across different study types in what 

5 information they provide. For example, because conditions can be controlled in laboratory 

6 studies, responses in such studies may be less variable and smaller differences may be easier to 

7 detect. However, the control conditions may limit the range of responses or incompletely reflect 

8 pollutant bioavailability, so they may not reflect responses that would occur in the natural 

9 environment. Alternatively, field data can provide important information for assessments of 

10 multiple stressors or where site-specific factors significantly influence exposure. They are also 

11 often useful for analyses of larger geographic scales and higher levels of biological organization. 

12 However, because most field study conditions can not be controlled, variability is expected to be 

13 higher and differences harder to detect. The presence of confounding factors can also make it 

14 difficult to attribute observed effects to specific stressors. 

15 In considering information from across multiple lines of evidence, our approach is to first 

16 integrate the evidence from both controlled and field-based studies and assess the coherence and 

17 consistency across the available evidence for each effect. We then consider the extent to which 

18 these identified effects should be considered adverse to the public welfare, relying largely on the 

19 paradigm used in the 2008 review and 2010 proposed reconsideration (e.g., 75 FR 3006). This 

20 paradigm recognizes that the significance to the public welfare of O3-induced effects on sensitive 

21 vegetation growing within the U.S. can vary depending on the nature of the effect, the intended 

22 use of the sensitive plants or ecosystems, and the types of environments in which the sensitive 

23 vegetation and ecosystems are located. Accordingly, any given O3-related effect on vegetation 

24 and ecosystems (e.g., biomass loss, crop yield loss, foliar injury) may be judged to have a 

25 different degree of impact on the public welfare depending, for example, on whether that effect 

26 occurs in a Class I area, a city park, or commercial cropland. Our approach takes this variation in 

27 the significance of O3-related vegetation effects into account in evaluating the currently available 

28 evidence with regard to the extent to which it calls into question the adequacy of the current 

29 standard and, as appropriate, indicates potential alternative standards that would be appropriate 

30 for the Administrator to consider. In the 2010 proposed reconsideration, the Administrator 

31 proposed to place the highest priority and significance on vegetation and ecosystem effects to 

32 sensitive species that are known to or are likely to occur in federally protected areas such as 

33 national parks and other Class I areas, or on lands set aside by States, Tribes and public interest 

34 groups to provide similar benefits to the public welfare (75 FR 3023/24). Our approach in this 

35 review considers whether newly available information would suggest any evolution to this 

36 paradigm, in particular in the context of considering associated ecosystem services. 
1-38

1 Finally, our approach continues to give great weight to the scientific evidence available in 

2 this and previous reviews indicating the relevance of cumulative, seasonal, concentration3 weighted exposures in inducing vegetation effects. Therefore, we continue to express exposures 

4 in terms of the W126 index, and continue to consider the important policy implications regarding 

5 selection of an appropriate exposure index for vegetation. Our approach also places primary 

6 emphasis on studies that evaluated plant response to exposures that were or can be described 

7 using such an index. The policy-relevant discussions in chapters 5 and 6 focus on vegetation 

8 effects evidence and exposure/risk information that can be associated with cumulative, seasonal 

9 peak-weighted exposures, where possible. Discussions pertaining to the adequacy of the current 

10 secondary standard will consider what cumulative seasonal exposures would be allowed under 

11 air quality that would just meet the current standard. 

12 1.3.2.2.2 Consideration of Exposure and Risk Estimates and Air Quality Analyses 

13 To put judgments about O3-related vegetation and ecosystem effects and services into a 

14 broader public welfare context, we consider national scale exposure and risk assessments 

15 described in the second draft WREA (U.S. EPA, 2014b). We particularly focused on the WREA 

16 quantitative risks related to three types of vegetation effects: foliar injury, biomass loss, and crop 

17 yield loss. These risks were assessed in a range of WREA analyses variously involving recent O3

18 monitoring data and/or national-scale model-adjusted air quality scenarios for the current 

19 secondary standard and, in some analyses, for a cumulative, seasonal W126 form at one or more 

20 levels (15, 11 and 7 ppm-hours). Our consideration of these WREA results provide insight into 

21 the extent to which the current or potential alternative standards would be expected to maintain 

22 distributions of cumulative, seasonal O3 exposures below those associated with adverse 

23 vegetation effects. 

24 With regard to quantitative O3 risks related to welfare effects and ecosystem services for 

25 foliar injury, we consider two main analyses in the WREA: a screening-level assessment of 214 

26 National Parks and a case study focused on three National Parks. In the screening-level 

27 assessment, O3 concentrations in national parks are assessed using criteria developed from a U.S. 

28 Forest Service nationwide dataset on foliar injury, ambient O3 concentrations (in terms of W126 

29 index) and soil moisture (which can influence susceptibility of vegetation to foliar injury). 

30 Additionally, we consider a case study for Class I areas (Great Smoky Mountain National Park, 

31 Rocky Mountain National Park, and Sequoia/Kings Canyon National Park). We consider results 

32 from this case study for three metrics: 1) percent of vegetation cover affected by foliar injury; 2) 

33 percent of trails affected by foliar injury; 3) estimates of species specific biomass loss within the 

34 case study area. We also consider qualitative analyses on ecosystem services effects for this 

35 endpoint. For example, the second draft WREA uses GIS mapping to illustrate where effects 
1-39

1 may be occurring and relates those areas to national scale statistics for recreational use and data 

2 on hiking trails, campgrounds and other park amenities that intersect with potentially affected 

3 areas. These are used to identify impacts on ecosystem services associated with recreation in 

4 national parks. We additionally consider analyses relating elevated O3 concentrations to 

5 increased vulnerability to fire risk regimes, insect attacks and impacts on hydrological cycles. 

6 With regard to risks related to biomass and crop yield loss, we consider WREA results 

7 based on exposure-response functions for tree and crop species that predict the growth or yield 

8 response of each species, based on the exposure patterns estimated within its growing region. To 

9 compare exposure-response across species, genotypes or experiments for which absolute 

10 response values may vary greatly, the second draft WREA instead uses estimates of relative 

11 biomass loss for trees or yield loss for crops. The WREA develops such estimates nationally and 

12 separately for more than 100 federally designated Class I areas. Additionally, we consider 

13 WREA-developed estimates of associated impacts on the agriculture and forestry sectors 

14 quantifying how O3 exposure to vegetation is estimated to affect the provision of timber and 

15 crops and carbon sequestration. We consider estimates for impacts related to tree biomass loss on 

16 ecosystem services such as pollution removal, carbon storage and sequestration in five urban 

17 case study areas. We consider biomass and crop yield loss estimates in light of advice from 

18 CASAC, as discussed in sections 5.3 and 5.4 below. 

19 In considering the amount of weight to place on the estimates of exposures and risks at or 

20 above specific W126 values described in the second draft WREA, our approach: 1) evaluates the 

21 weight of the scientific evidence concerning vegetation effects associated with those O3

22 exposures; 2) considers the importance, from a public welfare perspective, of the O3-induced 

23 effects on sensitive vegetation and associated ecosystem services that are known or anticipated to 

24 occur as a result of exposures at selected W126 values; and, 3) recognizes that predictions of 

25 effects associated with any given O3 exposure may be mitigated or exacerbated by actual 

26 conditions in the field (i.e., co-occurring modifying environmental and genetic factors). When 

27 considering analyses in the second draft WREA that involve discrete exposure levels or varying 

28 levels of severity of effects, our approach recognizes that welfare-relevant exposures are more 

29 appropriately viewed as a continuum with greater confidence and less uncertainty about the 

30 existence of welfare effects at higher O3 exposure concentrations and less confidence and greater 

31 uncertainty as one considers lower exposure concentrations. We recognize that there is no sharp 

32 breakpoint within the continuum ranging from concentrations at and above the level of the 

33 current secondary standard down to the lowest cumulative, seasonal W126 value assessed. In 

34 considering these results in this second draft PA, we consider both concerns about the potential 

35 for welfare effects and their severity with the increasing uncertainty associated and our 

36 understanding of the likelihood of such effects following exposures to lower O3 concentrations. 
1-40

1 1.3.2.2.3 Considerations Regarding Ambient O3 Concentration Estimates 

2 Attributable to Background Sources 

3 As noted above, our approach in this review utilizes recent advances in modeling 

4 techniques to estimate the contributions of U.S. anthropogenic, international anthropogenic, and 

5 natural sources to ambient O3 (discussed in detail in Chapter 2 of this document). Such model 

6 estimates can provide insights into the extent to which different types of emissions sources 

7 contribute to total ambient O3 concentrations. Our consideration of this issue in the current 

8 review is informed by the approaches taken in previous reviews, and by court decisions on 

9 subsequent litigation, as discussed in section 1.3.1.2.3 above. 

10 Further, in the 1996 proposal, O3 background concentrations were one of the factors the 

11 Administrator considered in selecting the SUM06 index as a form for an alternative secondary 

12 standard. This and other cumulative exposure indices under consideration were judged to be 

13 equally capable at estimating exposures relevant to vegetation, given the lack of evidence for a 

14 discernible threshold for vegetation effects in general (U.S. EPA 1996, p. 225), which might 

15 have provided a scientific basis for selecting among different cumulative exposure indices. At 

16 that time, the SUM06 metric was selected over the W126 metric because it focused on the 

17 policy-relevant (above background) portion of the total cumulative seasonal exposures reaching 

18 plants (62 FR 38856). At the conclusion of that review, the Administrator ultimately chose to set 

19 the secondary standard identical to the primary standard, including using the 8-hour average 

20 instead of a cumulative seasonal form (62 FR 38868). In the 2008 review, staff analyses 

21 concluded that the W126 index was more biologically-relevant based on the available science; 

22 staff additionally noted, based on then-available estimates of background, that this form was also 

23 not likely to be significantly impacted by background concentrations given the very low weight 

24 assigned to lower O3 concentrations by the W126 index (2007 SP, 7-22; 72 FR 37893). 

25 1.3.3 Organization of this Document 

26 Chapter 2 of this second draft PA provides an overview of the O3 ambient monitoring 

27 network and O3 air quality, including estimates of O3 concentrations attributable to background 

28 sources. The remaining chapters are organized into two main parts. Chapters 3 and 4 focus on the 

29 review of the primary O3 NAAQS while chapters 5 and 6 focus on the review of the secondary 

30 O3 NAAQS. Staff’s considerations and conclusions related to the current primary and secondary 

31 standards are discussed in chapters 3 and 5, respectively. Staff’s considerations and conclusions 

32 related to potential alternative primary and secondary standards are discussed in chapters 4 and 

33 6, respectively. Key uncertainties in the review and areas for future research and data collection 

34 are additionally identified in chapters 4 and 6 for the two types of standards. 
1-41

1 1.4 REFERENCES 

2 Frey, H.C. and Samet, J.M. (2012a) CASAC Review of the EPA’s Policy Assessment for the Review of the Ozone 

3 National Ambient Air Quality Standards (First External Review Draft – August 2012). EPA-CASAC-13-

4 003. November 26, 2012. 

5 Frey, H.C. and Samet, J.M. (2012b) CASAC Review of the EPA’s Health Risk and Exposure Assessment for Ozone 

6 (First External Review Draft - Updated August 2012) and Welfare Risk and Exposure Assessment for 

7 Ozone (First External Review Draft - Updated August 2012). EPA-CASAC-13-002. November 19, 2012 

8 Henderson, R. (2006) Letter from CASAC Chairman Rogene Henderson to EPA Administrator Stephen Johnson. 

9 October 24, 2006, EPA-CASAC-07-001.

10 Henderson, R. (2007) Letter from CASAC Chairman Rogene Henderson to EPA Administrator Stephen Johnson. 

11 March 26, 2007, EPA-CASAC-07-002.

12 Henderson, R. (2008) Letter from CASAC Chairman Rogene Henderson to EPA Administrator Stephen Johnson. 

13 April 7, 2008, EPA-CASAC-08-001.

14 Samet, J.M. (2011) Clean Air Scientific Advisory Committee (CASAC) Response to Charge Questions on the 

15 Reconsideration of the 2008 Ozone National Ambient Air Quality Standards. EPA-CASAC-11-004. 

16 March 30, 2011. Available online at: 

17 http://yosemite.epa.gov/sab/sabproduct.nsf/0/F08BEB48C1139E2A8525785E006909AC/$File/EPA18 CASAC-11-004-unsigned+.pdf 

19 U.S. Department of Health, Education, and Welfare (1970). Air Quality Criteria for Photochemical Oxidants. 

20 Washington, D.C.: National Air Pollution Control Administration; publication no. AP-63. Available from: 

21 NTIS, Springfield, VA; PB-190262/BA. 

22 U.S. EPA (U.S. Environmental Protection Agency). (1978). Air quality criteria for ozone and other photochemical 

23 oxidants [EPA Report]. (EPA/600/8-78/004). Washington, D.C.. 

24 U.S. EPA (U.S. Environmental Protection Agency). (1982). Air quality criteria document for ozone and other 

25 photochemical oxidants. Fed Reg 47: 11561. 

26 U.S. EPA (U.S. Environmental Protection Agency). (1986). Air quality criteria for ozone and other photochemical 

27 oxidants [EPA Report]. (EPA-600/8-84-020aF - EPA-600/8-84-020eF). Research Triangle Park, NC. 

28 http://www.ntis.gov/search/product.aspx?ABBR=PB87142949 

29 U.S. Environmental Protection Agency. (1989) Review of the National Ambient Air Quality Standards for Ozone: 

30 Policy Assessment of Scientific and Technical Information. OAQPS Staff Paper. Office of Air Quality 

31 Planning and Standards, Research Triangle Park, NC. 

32 U.S. EPA. (1996) Air Quality Criteria for Ozone and Related Photochemical Oxidants Volume I of III (Final, 1996). 

33 U.S. Environmental Protection Agency, Washington, D.C., EPA/600/AP-93/004aF (NTIS PB94173127). 

34 U.S. EPA. (2004) Air Quality Criteria for Particulate Matter (Final Report). U.S. Environmental Protection Agency, 

35 Washington, D.C., EPA 600/P-99/002aF-bF, 2004. 

36 U.S. Environmental Protection Agency. (2006). Air quality criteria for ozone and related photochemical oxidants 

37 [EPA Report]. (EPA/600/R-05/004AF). Research Triangle Park, NC. 

38 http://cfpub.epa.gov/ncea/cfm/recordisplay.cfm?deid=149923 
1-42

1 U.S. Environmental Protection Agency. (2007) Review of the National Ambient Air Quality Standards for Ozone: 

2 Policy Assessment of Scientific and Technical Information. OAQPS Staff Paper. Office of Air Quality 

3 Planning and Standards, Research Triangle Park, NC. EPA-452/R-07-007.

4 U.S. Environmental Protection Agency. (2011a). Integrated Review Plan for the O3 National Ambient Air Quality 

5 Standards (IRP) U. S. Environmental Protection Agency, National Center for Environmental Assessment 

6 Office of Research and Development and Office of Air Quality Planning and Standards Office of Air and 

7 Radiation, Research Triangle Park, North Carolina EPA. 452/R-11-006. 

8 U.S. Environmental Protection Agency. (2011b) Integrated Science Assessment for Ozone and Related 

9 Photochemical Oxidants: First External Review Draft, U.S. Environmental Protection Agency, 

10 Washington, D.C., EPA/600/R-10/076A. 

11 U.S. Environmental Protection Agency. (2011c) Integrated Science Assessment of Ozone and Related 

12 Photochemical Oxidants (Second External Review Draft). U.S. Environmental Protection Agency, 

13 Washington, D.C., EPA/600/R-10/076B. 

14 U.S. Environmental Protection Agency. (2011d) Policy Assessment for the Review of the Particulate Matter 

15 National Ambient Air Quality Standards. Office of Air Quality Planning and Standards, Research Triangle 

16 Park, NC. EPA 452/R-11-003.

17 U.S. Environmental Protection Agency. (2012a). Integrated Science Assessment for Ozone and Related 

18 Photochemical Oxidants: Third External Review Draft, U.S. Environmental Protection Agency, Research 

19 Triangle Park, NC. EPA/600/R-10/076C 

20 U.S. Environmental Protection Agency. (2012b). Health Risk and Exposure Assessment for Ozone, First External 

21 Review Draft, U.S. Environmental Protection Agency, Research Triangle Park, NC. EPA 452/P-12-001. 

22 U.S. Environmental Protection Agency. (2013). Integrated Science Assessment for Ozone and Related 

23 Photochemical Oxidants (Final Report). U.S. Environmental Protection Agency, Washington, D.C., 

24 EPA/600/R-10/076F, 2013 

25 U.S. Environmental Protection Agency. (2014a). Health Risk and Exposure Assessment for Ozone, Second External 

26 Review Draft. Office of Air Quality Planning and Standards, Research Triangle Park, NC. EPA-452/P-14-

27 004a. 

28 U.S. Environmental Protection Agency. (2014b). Welfare Risk and Exposure Assessment for Ozone, Second 

29 External Review Draft. Office of Air Quality Planning and Standards, Research Triangle Park, NC. EPA30 452/P-14-003a. 
2-1

1 2 O3 MONITORING AND AIR QUALITY 

2 This section provides overviews of ambient O3 monitoring in the U.S. (section 2.1); O3

3 precursor emissions and atmospheric chemistry (section 2.2); ambient O3 concentrations (section 

4 2.3); and available evidence and information related to background O3 (section 2.4). These 

5 issues are also discussed in detail in chapter 3 of the ISA (US EPA, 2013).

6 2.1 O3 MONITORING 

7 2.1.1 O3 Monitoring Network 

8 To monitor compliance with the NAAQS, state and local environmental agencies operate 

9 O3 monitoring sites at various locations, depending on the population of the area and typical peak 

O3 concentrations110 . In 2010, there were over 1,300 state, local, and tribal O3 monitors reporting 

11 concentrations to EPA. In areas for which O3 monitors are required, at least one site must be 

12 designed to record the maximum concentration for that particular metropolitan area. Since O3

13 concentrations are usually significantly lower in the colder months of the year, O3 is required to 

be monitored only during the O3 monitoring season, which varies by state.214

15 Figure 2-1 shows the locations of the U.S. ambient O3 monitoring sites reporting data to 

16 EPA at any time during the 2006-2010 period. The gray dots which make up over 80% of the O3

17 monitoring network are “State and Local Monitoring Stations” (SLAMS) monitors, which are 

18 operated by state and local governments to meet regulatory requirements and provide air quality 

19 information to public health agencies. Thus, the SLAMS monitoring sites are largely focused on 

20 urban areas. The blue dots highlight two important subsets of monitoring sites within the 

21 SLAMS network: the “National Core” (NCore) multi-pollutant monitoring network and the 

22 “Photochemical Assessment Monitoring Stations” (PAMS) network. 

23 While the existing U.S. O3 monitoring network has a largely urban focus, to address 

24 ecosystem impacts of O3 such as biomass loss and foliar injury, it is equally important to focus 

25 on O3 monitoring in rural areas. The green dots in Figure 2-1 represent the Clean Air Status and 

26 Trends Network (CASTNET) monitors which are located in rural areas. There were about 80 

27 CASTNET sites operating in 2010, with sites in the eastern U.S. being operated by EPA and 

28 sites in the western U.S. being operated by the National Park Service (NPS). Finally, the black 

29 dots represent “Special Purpose Monitoring Stations” (SPMS), which include about 20 rural 

30 monitors as part of the “Portable O3 Monitoring System” (POMS) network operated by the NPS. 

 

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

 The required O3 monitoring seasons for each state are listed in Table D-3 of Appendix D to 40 CFR Part 58. 
2-2

1 Between the CASTNET, NCore, and POMS networks, there were about 120 rural O3 monitoring 

2 sites operating in the U.S. in 2010. 

3

4

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

6 2006-2010 period.

7 2.1.2 Recent O3 Monitoring Data and Trends 

8 To determine whether or not the O3 NAAQS has been met at an ambient monitoring site, 

9 a statistic commonly referred to as a “design value” must be calculated based on 3 consecutive 

10 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 11 highest daily maximum 8-hour O3 concentration in parts 

12 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.313 In 

14 counties or other geographic areas with multiple monitors, the area-wide design value is defined 

15 as the design value at the highest individual monitoring site, and the area is said to have met the 

16 NAAQS if all monitors in the area are meeting the NAAQS. 

Figure 2-2 shows the trend in the annual 4th 17 highest daily maximum 8-hour O3

18 concentrations in ppb based on 933 “trends” sites with complete data records over the 2000 to 

 

3

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. 
2-3

1 2012 period. The center line in this figure represents the median value across the trends sites, 

while the dashed lines represent the 25th and 75th 2 percentiles, and the bottom and top lines 

represent the 10th and 90th 3 percentiles. Figures 2-3 and 2-4 show maps of the O3 design values 

4 (ppb) at all U.S. monitoring sites for the 2009-2011 and 2010-2012 periods, respectively. The 

trend shows that the annual 4th 5 highest daily maximum values decreased for the vast majority of 

6 monitoring sites in the U.S. between 2000 and 2009. The decreasing trend is especially sharp 

7 from 2002 to 2004, when EPA implemented the “NOx SIP Call”, a program designed to reduce 

8 summertime emissions of NOx in the eastern U.S. 

The trends also show a modest increase in the 4th 9 highest daily maximum values from 

10 2009 to 2012. This is reflected in the design value maps, which show an increase in the number 

11 of monitors violating the existing O3 standard in 2010-2012 relative to 2009-2011. Meteorology 

12 played an important role in these trends. O3 concentrations tend to be higher on days with hot 

13 and stagnant conditions and lower on days with cool or wet conditions. According to the 

14 National Oceanic and Atmospheric Administration’s National Climactic Data Center (NOAA15 NCDC), the summer of 2009 was cooler and wetter than average over most of the eastern U.S., 

16 while conversely, the summers, of 2010, 2011, and 2012 were all much warmer than average. In 

17 particular, the central and eastern U.S. experienced a 2-week period of record-breaking heat in 

18 late June and early July of 2012, which contributed to hundreds of violations of the existing O3

19 standard. In contrast, the most recent climatological information available from NOAA-NCDC 

20 (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 21

22 highest daily maximum O3 concentrations to continue in 2013. 
2-4

1

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

3 ppb, 2000 to 2012. 

4

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

1

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

3 In addition, EPA focused our analyses of welfare and ecosystem effects on a W126 O3

exposure metric in this review. The W126 metric44 is a seasonal aggregate of daytime (8:00 AM 

5 to 8:00 PM) hourly O3 concentrations, designed to measure the cumulative effects of O3

6 exposure on vulnerable plant and tree species, with units in parts per million-hours (ppm-hrs). 

7 The W126 metric uses a logistic weighting function to place less emphasis on exposure to low 

8 hourly O3 concentrations and more emphasis on exposure to high hourly O3 concentrations 

9 (Lefohn et al, 1988). 

10 Figure 2-5 shows the trend in annual W126 concentrations in ppm-hrs based on 933 

11 “trends” sites with complete data records over the 2000 to 2012 period. Figures 2-6 and 2-7 

12 show maps of the 3-year average annual W126 concentrations in ppm-hrs at all U.S. monitoring 

13 sites for the 2009-2011 and 2010-2012 periods, respectively. The general patterns seen in these 

14 figures are similar to those seen in the design value metric for the existing standard. 

15

 

4

 Details on the procedure used to calculate the W126 metric are provided in Chapter 4 of the welfare Risk and 

Exposure Assessment. 
2-6

1

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

3

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

1

2

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

4 2.2 EMISSIONS AND ATMOSPHERIC CHEMISTRY 

5 O3 is formed by photochemical reactions of precursor gases and is not directly emitted 

6 from specific sources. In the stratosphere, O3 occurs naturally and provides protection against 

7 harmful solar ultraviolet radiation. In the troposphere, near ground level, O3 forms through 

8 atmospheric reactions involving two main classes of precursor pollutants: non-methane volatile 

9 organic compounds (NMVOCs) and nitrogen oxides (NOX). Carbon monoxide (CO) and 

10 methane (CH4) are also important for O3 formation over longer time periods (US EPA, 2013, 

11 section 3.2.2). 

12 Emissions of O3 precursor compounds can be divided into anthropogenic and natural 

13 source categories, with natural sources further divided into biogenic emissions (from vegetation, 

14 microbes, and animals) and abiotic emissions (from biomass burning, lightning, and geogenic 

15 sources). Anthropogenic sources, including mobile sources and power plants, account for the 

16 majority of NOx and CO emissions. Anthropogenic sources are also important for NMVOC 

17 emissions, though in some locations and at certain times of the year (e.g., southern states during 

18 summer) the majority of NMVOC emissions come from vegetation (US EPA, 2013, section 

19 3.2.1). In practice, the distinction between natural and anthropogenic sources is often unclear, as 

20 human activities directly or indirectly affect emissions from what would have been considered 
2-8

1 natural sources during the preindustrial era. Thus, emissions from plants, animals, and wildfires 

2 could be considered either natural or anthropogenic, depending on whether emissions result from 

3 agricultural practices, forest management practices, lightning strikes, or other types of events 

4 (US EPA, 2013, sections 3.2 and 3.7.1). 

5 Rather than varying directly with emissions of its precursors, O3 changes in a nonlinear 

6 fashion with the concentrations of its precursors. NOx emissions lead to both the formation and 

7 destruction of O3, depending on the local quantities of NOx, NMVOC, radicals, and sunlight. In 

8 areas dominated by fresh emissions of NOx, these radicals are removed, which lowers the O3

9 formation rate. In addition, the scavenging of O3 by reaction with NO is called “titration” and is 

10 often found in downtown metropolitan areas, especially near busy streets and roads, and in 

11 power plant plumes. This short-lived titration results in localized areas in which O3

12 concentrations are suppressed compared to surrounding areas, but which contain NO2 that 

13 contributes to O3 formation later and further downwind. Consequently, O3 response to 

14 reductions in NOx emissions is complex and may include O3 decreases at some times and 

15 locations and increases of O3 at other times and locations. In areas with low NOx to VOC ratios, 

16 such as those found in remote continental areas and rural and suburban areas downwind of urban 

17 centers, O3 production typically varies directly with NOx concentrations (e.g. increases with 

18 increasing NOx emissions). 

19 At some times and in some locations, reductions in O3 precursors may also yield 

20 reductions in ambient air pollutants other than O3. For example, given that NOx emissions 

21 contribute to ambient NO2 (i.e., both because NO2 is a component of NOx emissions and because 

22 NO can convert rapidly to NO2), reductions in directly emitted NOx will also result in reductions 

23 in ambient NO2. In addition, NOx and VOCs can contribute to secondary formation of PM2.5 

24 constituents. NOx can act as both a direct precursor to NH4NO3, and can affect the formation of 

other PM2.5 constituents because it adds to the oxidative capacity of the atmosphere525 . The effects 

26 of reducing NOx emissions on ambient PM2.5 concentrations can vary in time and space, with the 

27 largest reductions in ambient PM2.5 likely occurring at times when and in locations where 

28 concentrations of NH4NO3 are highest. This is usually during the cooler times of the year (e.g. 

29 April-November) and in some areas of California, Salt Lake City, The Great Lakes States, and 

the Northeast corridor between Baltimore and New York City (Carlton et al, 2010)630 .

31 The formation of O3 from precursor emissions is also affected by the intensity and 

32 spectral distribution of sunlight and atmospheric mixing. Major episodes of high ground-level 

33 O3 concentrations in the eastern United States are associated with slow-moving high pressure 

 

5

 Across North America, approximately 7% of summertime PM2.5 mass is estimated to result from anthropogenic 

NOx emissions and up to 0.5 ug/m3 of secondary organic aerosol is estimated to form from NOx emissions (Carlton 

et al, 2010). 

6

 In these locations, NH4NO3 contributes more than 30% to average PM2.5 concentrations. 
2-9

1 systems. High pressure systems during the warmer seasons are associated with the sinking of 

2 air, resulting in warm, generally cloudless skies, with light winds. The sinking of air results in 

3 the development of stable conditions near the surface which inhibit or reduce the vertical mixing 

4 of O3 precursors. The combination of inhibited vertical mixing and light winds minimizes the 

5 dispersal of pollutants emitted in urban areas, allowing their concentrations to build up. In 

6 addition, in some parts of the United States (e.g., in Los Angeles), mountain barriers limit mixing 

7 and result in a higher frequency and duration of days with high O3 concentrations. 

8 Photochemical activity involving precursors is enhanced during warmer seasons because of the 

9 availability of sunlight and higher temperatures (US EPA, 2013, section 3.2). 

10 O3 concentrations in a region are affected both by local formation and by transport of O3

11 and its precursors from upwind areas. O3 transport occurs on many spatial scales including local 

12 transport between cities, regional transport over large regions of the U.S. and international/long13 range transport. In addition, O3 can be transferred into the troposphere from the stratosphere, 

14 which is rich in O3, through stratosphere-troposphere exchange (STE). These inversions or 

15 “folds” usually occur behind cold fronts, bringing stratospheric air with them and typically affect 

16 O3 concentrations in high elevation areas (e.g. > 1500 m) more than areas at low elevations (U.S. 

17 EPA, 2012, section 3.4.1.1). The role of long-range transport of ozone and other elements of 

18 ozone background is discussed in more detail in Section 2.4, 

19 2.3 AIR QUALITY CONCENTRATIONS 

20 Because O3 is a secondary pollutant formed in the atmosphere from precursor emissions, 

21 concentrations are generally more regionally homogeneous than concentrations of primary 

22 pollutants emitted directly from stationary and mobile sources (US EPA, 2013, section 3.6.2.1). 

23 However, variation in local emissions characteristics, meteorological conditions, and topography 

24 can result in daily and seasonal temporal variability in ambient O3 concentrations, as well as 

25 local and national-scale spatial variability. 

26 Temporal variation in ambient O3 concentrations results largely from daily and seasonal 

27 patterns in sunlight, precursor emissions, atmospheric stability, wind direction, and temperature 

28 (US EPA, 2013, section 3.7.5). On average, ambient O3 concentrations follow well-recognized 

29 daily and seasonal patterns, particularly in urban areas. Specifically, daily maximum O3

30 concentrations in urban areas tend to occur in mid-afternoon, with more pronounced peaks in the 

31 warm months of the O3 season than in the colder months (US EPA, 2013, Figures 3-54, 3-156 to 

32 3-157). Rural sites also followed this general pattern, though it is less pronounced in colder 

33 months (US EPA, 2013, Figure 3-55). With regard to day-to-day variability, median maximum 

34 daily average 8-hour (MDA8) O3 concentrations in U.S. cities in 2007-2009 were approximately 
2-10

1 47 ppb, with typical ranges between 35 to 60 ppb and the highest MDA8 concentrations above 

2 100 ppb in several U.S. cities (as noted further below). 

3 In addition to temporal variability, there is considerable spatial variability in ambient O3

4 concentrations within cities and across different cities in the United States. With regard to 

5 spatial variability within a city, local emissions characteristics, geography, and topography can 

6 have important impacts. For example, as noted above, fresh NO emissions from motor vehicles 

7 titrate O3 present in the urban background air, resulting in an O3 gradient around roadways with 

8 O3 concentrations increasing as distance from the road increases (US EPA, 2013, section 

9 3.6.2.1). In comparing urban areas, the ISA notes that measured O3 concentrations are relatively 

10 uniform and well-correlated across some cities (e.g., Atlanta) while they are more variable in 

11 others (e.g., Los Angeles) (US EPA, 2013, section 3.6.2.1 and Figures 3-28 to 3-36). 

12 With regard to spatial variability across cities, when the ISA evaluated the distributions 

13 of 8-hour O3 concentrations for the years 2007 to 2009 in 20 cities, the highest concentrations 

14 were reported in Los Angeles, with high concentrations also reported in several eastern and 

15 southern cities. The maximum recorded MDA8 was 137 ppb in Los Angeles, and was near or 

16 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 17 percentile of the distribution of 

MDA8 concentrations7

, with Los Angeles recording the highest 98th 18 percentile concentration (91 

ppb) and many eastern and southern cities reporting 98th 19 percentile concentrations near or above 

75 ppb. In contrast, somewhat lower 98th 20 percentile O3 concentrations were recorded in cities in 

21 the western United States outside of California (US EPA, 2013, Table 3-10).

22 Although rural monitoring sites tend to be less directly affected by anthropogenic 

23 pollution sources than urban sites, rural sites can be affected by transport of O3 or O3 precursors 

24 from upwind urban areas and by local anthropogenic sources such as motor vehicles, power 

25 generation, biomass combustion, or oil and gas operations (US EPA, 2013, section 3.6.2.2). In 

26 addition, O3 tends to persist longer in rural than in urban areas due to lower rates of chemical 

27 scavenging in non-urban environments. At higher elevations, increased O3 concentrations can 

28 also result from stratospheric intrusions (US EPA, 2013, sections 3.4, 3.6.2.2). As a result, O3

29 concentrations measured in some rural sites can be higher than those measured in nearby urban 

30 areas (US EPA, 2013, section 3.6.2.2), and the ISA concludes that cumulative exposures for 

31 humans and vegetation in rural areas can be substantial, often higher than cumulative exposures 

32 in urban areas (US EPA, 2013, section 3.7.5). 

 

7

 Table 3-10 in the ISA analyzes the warm season. Therefore, the 98th percentile values would be an approximation 

of the 4th highest value. 
2-11

1 2.4 BACKGROUND O3

2 Generically, background O3 can originate from natural sources of O3 and O3 precursors, 

3 as well as from manmade international emissions of O3 precursors. Natural sources of O3

4 precursor emissions such as wildfires, lightning, and vegetation can lead to O3 formation by 

5 chemical reactions with other natural sources. Another important natural component of 

6 background is O3 that is naturally formed in the stratosphere through interactions of ultraviolet 

7 light with molecular oxygen. Stratospheric O3 can mix down to the surface at high 

8 concentrations in discrete events called intrusions, especially at higher-altitude locations. The 

9 manmade portion of the background includes any O3 formed due to anthropogenic sources of O3

10 precursors emitted far away from the local area (e.g., international emissions). Finally, both 

11 biogenic and international anthropogenic emissions of methane, which can be chemically 

12 converted to O3 over relatively long time scales, can also contribute to global background O3

13 levels. 

14 As indicated in the first draft policy assessment (US EPA, 2012, sections 1.3.4 and 3), 

15 EPA has updated several aspects of our methodology for estimating the change in health risk and 

16 exposure that would result from a revision to the O3 NAAQS. First, risk estimates are now based 

17 on total O3 concentrations, as opposed to previous reviews which only considered risk above 

18 background levels. Second, EPA is now using air quality models to estimate the spatial patterns 

19 of O3 that would result from attaining various levels of the NAAQS, as opposed to a quadratic 

20 rollback approach that required the estimation of a background “floor” beyond which the 

21 rollback would not take place. Both of these revisions have had the indirect effect of reducing 

22 the need for estimates of background O3 levels as part of the O3 risk and exposure assessment 

23 (REA). Regardless, EPA expects that a well-founded understanding of the fractional 

24 contribution of background sources and processes to surface O3 levels will be valuable towards 

25 informing policy decisions about the O3 NAAQS. Accordingly, in this section, we briefly 

26 summarize existing results on background O3 from the ISA (US EPA, 2013, section 3.4) as 

27 supplemented by additional EPA modeling recently conducted for a 2007 base year. The 

28 summary will focus on national estimates of: 1) seasonal mean background O3 concentrations for 

29 three specific definitions of background O3, 2) the relative proportion of background O3 to total 

30 O3 for the same three definitions from a seasonal mean perspective, 3) the distributions of 

31 background O3 within a seasonal mean, 4) the fractional background O3 in the 12 REA urban 

32 case study areas, 5) the relative proportion of background O3 concentrations to total ozone from a 

33 W126 perspective, and 6) the relative roles of different components of background O3.

34 The definition of background O3 can vary depending upon context, but it generally refers 

35 to O3 that is formed by sources or processes that cannot be influenced by actions within the 

36 jurisdiction of concern. In the first draft policy assessment document (US EPA, 2012), EPA 
2-12

1 presented three specific definitions of background O3: natural background, North American 

2 background, and United States background. Natural background (NB) was the narrowest 

3 definition of background, and it was defined as the O3 that would exist in the absence of any 

4 manmade O3 precursor emissions. The other two previously established definitions of 

5 background presume that the U.S. has little influence over anthropogenic emissions outside our 

6 continental or domestic borders. North American background (NAB) is defined as that O3 that 

7 would exist in the absence of any manmade O3 precursor emissions from North America. U.S. 

8 background (USB) is defined as that O3 that would exist in the absence of any manmade 

9 emissions inside the United States. Each of these three definitions of background O3 requires 

10 photochemical modeling simulations to estimate what the residual O3 concentrations would be 

11 were the various anthropogenic emissions to be removed. Previous modeling studies have 

12 estimated what background levels would be in the absence of certain sets of emissions by simply 

13 assessing the remaining O3 in a simulation in which certain emissions were removed (Zhang et 

14 al. (2011), Emery et al. (2012)). This basic approach is often referred to as “zero-out” modeling 

15 or “emissions perturbation” modeling. 

16 While the zero-out approach has traditionally been used to estimate background O3

17 levels, the methodology has some acknowledged limitations. First, from a policy perspective, 

18 the hypothetical and unrealizable zero manmade emissions scenarios have limited application. 

19 Secondly, the assumption that background O3 is what is left after specific emissions have been 

20 removed within the model simulation can be misleading in locations where O3 chemistry is 

21 highly non-linear. Depending upon the local composition of O3 precursors, NOx emissions 

22 reductions can either increase or decrease O3 levels in the immediate vicinity of those reductions. 

23 For those specific urban areas in which NOx titration of O3 can be significant, zero-out modeling 

24 can result in inflated estimates of background O3 when these NOx emissions are completely and 

25 unrealistically removed. Paradoxically, in certain times and locations in a zero-out scenario 

26 there can be more background O3 than actual O3 within the model. A separate modeling 

27 technique circumvents these limitations by apportioning the total O3 within the model to its 

28 contributing source terms. This basic approach, referred to as “source apportionment” modeling, 

29 has been described and evaluated in the peer-reviewed literature (Dunker et al., 2002; Kemball30 Cook et al., 2009). While source apportionment modeling has not been previously used in the 

31 context of estimating background ozone levels as part of an ozone NAAQS review, it has 

32 frequently been used in other regulatory settings to estimate the “contribution” to ozone of 

33 certain sets of emissions (EPA 2005, EPA 2011). The source apportionment technique provides 

34 a means of estimating the contributions of user-identified source categories to ozone formation in 

35 a single model simulation. This is achieved by using multiple tracer species to track the fate of 

36 ozone precursor emissions (VOC and NOx) and the ozone formation caused by these emissions. 
2-13

1 The methodology is designed so that all ozone and precursor concentrations are attributed to the 

2 selected source categories at all times without perturbing the inherent chemistry. EPA recently 

3 completed updated zero-out and source apportionment modeling for a 2007 base year to 

4 supplement our characterization of background O3 over the U.S. Prior to using model 

5 simulations to estimate background O3 levels over the U.S., EPA confirmed that the modeling 

6 was able to reproduce historical O3 levels and that there was limited correlation between model 

7 errors and the background estimates. The key findings from the updated modeling are described 

8 below; a more detailed description of the modeling is provided in Appendix A. 

9 2.4.1 Seasonal Mean Background O3 in the U.S. 

10 The ISA (US EPA 2013, section 3.4) previously established that background 

11 concentrations vary spatially and temporally and that simulated mean background concentrations 

12 are highest at high-elevation sites within the western U.S. Background levels typically are 

13 greatest over the U.S. in the spring and early summer. Figure 2-8 displays the spatial patterns of 

seasonal mean814 natural background O3 as estimated by a 2007 zero-out scenario. This figure 

15 shows the average daily maximum 8-hour O3 concentration (MDA8) that would exist in the 

16 absence of any anthropogenic O3 precursor emissions at monitor locations. Seasonal mean NB 

17 levels range from approximately 15-35 ppb with the highest values at higher-elevation sites in 

18 the western U.S. The median value over these locations is 24.2 ppb, and more than 50 percent of 

19 the locations have natural background levels of 20-25 ppb. The highest modeled estimate of 

20 seasonal average, natural background, MDA8 O3 is 34.3 ppb at the high-elevation CASTNET 

21 site (Gothic) in Gunnison County, CO. Natural background levels are higher at these high22 elevation locations primarily because natural stratospheric O3 impacts and international transport 

23 impacts increase with altitude (where O3 lifetimes are longer). 

 

8

 The recent EPA modeling focused on the period from April through October. Seasonal means are computed over 

those seven months. 
2-14

1

2 Figure 2-8. Map of 2007 CMAQ-estimated seasonal mean natural background O3 levels 

3 from zero-out modeling.

4 Figures 2-9 and 2-10 show the same information for the NAB and USB scenarios. In 

5 these model runs, all anthropogenic O3 precursor emissions were removed from the U.S., 

6 Canada, and Mexico portions of the modeling domain (NAB scenario) and then only from the 

7 U.S. (USB scenario). The figures show that there is not a large difference between the NAB and 

8 USB scenarios. Seasonal mean NAB and USB O3 levels range from 25-50 ppb, with the most 

9 frequent values estimated in the 30-35 ppb bin. The median seasonal mean background levels 

10 are 31.5 and 32.7 ppb (NAB and USB, respectively). Again, the highest levels of seasonal mean 

11 background are predicted over the intermountain western U.S. Locations with NAB and USB 

12 concentrations greater than 40 ppb are confined to Colorado, Nevada, Utah, Wyoming, northern 

13 Arizona, eastern California, and parts of New Mexico. The 2007 EPA modeling suggests that 

14 seasonal mean USB concentrations are on average 1-3 ppb higher than NAB background. These 

15 results were similar to those reported by Wang et al. (2009). From a seasonal mean perspective, 

16 background levels are typically well-below the NAAQS thresholds. 
2-15

1

2 Figure 2-9. Map of 2007 CMAQ-estimated seasonal mean North American background 

3 O3 levels from zero-out modeling. 

4

5 Figure 2-10. Map of 2007 CMAQ-estimated seasonal mean United States background O3

6 levels from zero-out modeling. 

7 2.4.2 Seasonal Mean Background O3 in the U.S. as a Proportion of Total O3

8 Another informative way to assess the importance of background as part of seasonal 

9 mean O3 levels across the U.S. is to consider the fractional contribution of NB, NAB, and USB 

10 to total modeled O3 at each monitoring location. Considering the proportional role of 
2-16

1 background allows for an initial assessment of the relative importance of background and non2 background sources. Figures 2-11 and 2-12 show the percent contribution of U.S. anthropogenic 

3 sources to total O3 using the metric of the seasonal mean MDA8 O3 concentrations as estimated 

4 by both the zero-out and source apportionment modeling methodologies. Recall that the terms 

5 NB, NAB, and USB are explicitly linked to the zero-out modeling approach. For comparison 

6 sake, in Figure 2-12 we are extending the definition of USB to also include the source 

7 apportionment model estimates of the O3 that is attributable to sources other than U.S. 

8 anthropogenic emissions. To preserve the original definition of USB, this second term will be 

9 hereafter referred to as “apportionment-based USB”. As noted earlier, the advantage of the 

10 source apportionment modeling is that all of the modeled O3 is attributed to various source terms 

11 and thus this approach is not affected by the confounding occurrences of background O3 values 

12 exceeding the base O3 values as can happen in the zero-out modeling (i.e., background 

13 proportions > 100%). Consequently, one would expect the fractional background levels to be 

14 lower in the source apportionment methodology as a result of removing this artifact. 

15 When averaged over all sites, O3 from sources other than U.S. anthropogenic emissions 

16 is estimated to comprise 66 (zero-out) and 59 (source apportionment) percent of the total 

17 seasonal O3 mean. The spatial patterns of apportionment-based USB are similar across the two 

18 modeling exercises. Background O3 is a relatively larger percentage (e.g., 70-80%) of the total 

19 seasonal mean O3 in locations within the intermountain western U.S. and along the U.S. border. 

20 In locations where O3 levels are generally higher, like California and the eastern U.S. the 

21 seasonal mean background fractions are relatively smaller (e.g., 40-60%). The additional 2007 

22 modeling confirms that background ozone, while generally not approaching levels of the ozone 

23 standard, can comprise a considerable fraction of total seasonal mean ozone across the U.S. 

24 2.4.3 Daily Distributions of Background O3 within the Seasonal Mean 

25 As a first-order understanding, it is valuable to be able to characterize seasonal mean 

26 levels of background O3. However, it is well established that background levels can vary 

27 substantially from day-to-day. From an implementation perspective, the values of background 

28 O3 on possible exceedance days are a more meaningful consideration. The first draft policy 

29 assessment (US EPA, 2012) considered this issue in detail, via summaries of the existing 2006 

30 zero-out modeling (Henderson et al., 2012), and concluded that “results suggest that background 

31 concentrations on the days with the highest total O3 concentrations are not dramatically higher 

32 than typical seasonal average background concentrations.” Based on this finding, EPA 

33 determined that “anthropogenic sources within the U.S. are largely responsible for 4th highest 8-

34 hour daily maximum O3 concentrations.” The recent EPA modeling using a 2007 base year and 

35 the two distinct modeling methodologies supports this finding from the previous 2006-based 
2-17

1 modeling analyses. That is, the highest modeled O3 site-days tend to have background O3 levels 

2 similar to mid-range O3 days. Figure 2-13 and 2-14 show the distribution of daily MDA8 

3 apportionment-based USB levels (absolute magnitudes and relative fractions, respectively) from 

the CAMx source apportionment simulation94 . Again, the 2007 modeling shows that the days 

5 with highest O3 levels have similar distributions (i.e., means, inter-quartile ranges) of 

6 background levels as days with lower values down to approximately 40 ppb. As a result, the 

7 proportion of total O3 that has background origins is smaller on high O3 days (e.g., days > 70 

8 ppb) than the more common lower O3 days that drive seasonal means. This helps put the results 

9 from section 2.4.2 into better context. For example, for site-days in which base O3 is between 

10 70-75 ppb, the source apportionment modeling estimates that approximately 37 percent of those 

11 O3 levels originate from sources other than U.S. anthropogenic emissions (i.e., apportionment12 based USB). Figure 2-14 also indicates that there are cases in which the model predicts much 

13 larger background proportions, as shown by the upper outliers in the figure. These infrequent 

14 episodes usually occur in relation to a specific event, and occur more often in specific 

15 geographical locations, such as at high elevations or wildfire prone areas during the local dry 

16 season. 

17 It should be noted here that EPA has policies for treatment of air quality monitoring data 

18 affected by these types of events. EPA’s exceptional events policy allows exclusion of certain 

19 air quality monitoring data from regulatory determinations if a State adequately demonstrates 

20 that an exceptional event has caused the exceedance or violation of a NAAQS. In addition, 

21 Section 179B of the CAA also provides for treatment of air quality data from international 

22 transport when an exceedance or violation of a NAAQS would not have occurred but for 

23 emissions emanating from outside of the United States. From an overarching perspective, the 

24 Clean Air Act requires the NAAQS to be set at a level requisite to protect public health and 

25 welfare. Case law makes it clear that attainability and technical feasibility are not relevant 

26 considerations in the setting of a NAAQS. In previous reviews, EPA has assessed the proximity 

27 of ozone concentrations to peak background levels only as a secondary consideration between 

28 potential threshold levels where health and welfare was determined to have been protected.

29

 

9

 Similar plots from the zero-out modeling for natural background, North American background, and U.S. 

background are provided in Appendix A. 
2-18

1

2 Figure 2-11. Map of U.S. background percent contribution to seasonal mean O3 based on 

3 2007 CMAQ zero-out modeling. 

4

5

6 Figure 2-12. Map of apportionment-based U.S. background percent contribution to 

7 seasonal mean O3 based on 2007 CAMx source apportionment modeling. 
2-19

1

2 Figure 2-13. Distributions of absolute estimates of apportionment-based U.S. Background 

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

4 apportionment simulation. 

5

6

7 Figure 2-14. Distributions of the relative proportion of apportionment-based U.S. 

8 Background to total O3 (all site-days), binned by modeled MDA8 from the 

9 2007 source apportionment simulation. 
2-20

1

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

3 As noted in the introduction, the current approach to estimating changes in risk across 

4 various possible levels of the O3 NAAQS no longer requires a quantitative estimate of 

5 background O3 levels. However, EPA expects to be asked the question: “how much of the total 

6 risk is due to background sources.” This section of the policy assessment presents estimates of 

7 the overall fraction of O3 that is estimated to result from background sources or processes in each 

8 of the 12 urban case study areas considered in the epidemiological-based risk assessment of the 

9 REA (US EPA 2014, Chapter 7). The results are based on the recent EPA 2007 source 

10 apportionment modeling. Table 2-1 summarizes the estimated fractional contributions of sources 

11 other than U.S. anthropogenic emissions (i.e., apportionment-based USB) to total seasonal mean 

12 MDA8 O3 in each of the 12 urban case study areas. The table shows that the fractional 

13 contributions from sources other than anthropogenic emissions within the U.S. can range from 

14 43 to 66 percent across these 12 urban areas. These fractions are consistent with the national 

15 ratios summarized in section 2.4.2, although the fractions of background are generally smaller at 

16 urban sites than at rural sites. 

17 As shown in section 2.4.3, the fractional contributions from background are smaller on 

18 days with high modeled O3 (i.e., days that may exceed the level of the NAAQS). Table 2-2

19 provides the fractional contributions from these apportionment-based USB sources, only 

20 considering days in which base model MDA8 O3 was greater than 60 ppb. As expected, the 

21 fractional background contributions are smaller, ranging from 31 to 55 percent. 

22 Rather than taking the fractions of the seasonal means (as in Table 2-1), Table 2-3

23 displays the mean and median daily MDA8 background fractions. These metrics may be more 

24 appropriate for application to health studies. The fractional contributions to backgrounds 

25 calculated via this approach are very similar to the Table 2-1 calculations. Although EPA 

26 expects the source apportionment results to provide a more realistic estimate of fractional 

27 background values, for completeness, we also provide USB fractions based on the zero-out 

28 modeling for the 12 cities (see Table 2-4). The results are similar to the source apportionment 

29 findings (Table 2-1), though the zero-out technique provides slightly higher background 

30 proportions, as expected. It should be noted that all fractional contributions are based on recent 

31 conditions from 2007. These fractional contributions would be expected to change as 

32 anthropogenic emissions and O3 levels are lowered. Based on the source apportionment 

33 modeling for these 12 areas, there is evidence that background levels comprise a non-negligible 

34 fraction of the total ozone observed within these locations. However, for site-days in which 

35 model MDA8 ozone exceeds 60 ppb, ozone formed from U.S. anthropogenic emissions comprise 
2-21

1 an even larger fraction of the total ozone in 11 of the 12 areas. The major metropolitan areas in 

2 the eastern U.S. (e.g., Atlanta, New York City, Philadelphia) are less influenced by background 

3 sources than a higher-elevation, western U.S., location like Denver. Even in Denver, though, 

4 U.S. anthropogenic emissions have a large influence on total ozone (45 percent). 

5 Table 2-1. Seasonal mean MDA8 O3 (ppb), seasonal mean apportionment-based USB 

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

7 total O3 (all site-days) in the 12 REA urban case study areas (%).

8

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

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

11 total O3 (site-days > 60 ppb) in the 12 REA urban study areas (%). 

12

13 Table 2-3. Fractional contribution of apportionment-based USB in the 12 REA urban 

14 study areas (%), using the means and medians of daily MDA8 fractions 

15 (instead of fractions of seasonal means). 

16

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

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
2-22

1 Table 2-4. Seasonal mean MDA8 O3 (ppb), seasonal mean USB contribution (ppb), and 

2 fractional USB contribution to total O3 (all site-days) in the 12 REA urban case 

3 study areas (%). 

4

5 2.4.5 Influence of Background O3 on W126 levels 

6 EPA also conducted a limited assessment of the impacts of background O3 sources on the 

7 W126 metric. The W126 metric (LeFohn et al., 1988) is a cumulative peak-weighted index 

8 designed to estimate longer-term effects of daytime ozone levels on sensitive vegetation and 

9 ecosystems. EPA used the 2007 zero-out modeling to assess NB, NAB, and USB influences at 

10 four sample locations: Atlanta GA, Denver CO, Farmington NM, and Riverside CA. As shown 

11 in Figure 2-7, each of the four analyses locations had relatively high observed values of W126 in 

12 2010-2012. 

13 As discussed above, in previous EPA reviews of the O3 NAAQS, the influence of 

14 background ozone was estimated according to a counterfactual assumption. Background O3 was 

15 defined as the ozone that would exist in the absence of a particular set of emissions (e.g., NAB is 

16 the ozone that would exist if there were no anthropogenic emissions in North America). In the 

17 current review, EPA is supplementing the counterfactual assessment with analyses that estimate 

18 the fraction of the existing ozone that is due to background sources. This has important 

19 ramifications for assessing the influence of background on W126 concentrations, because of the 

20 non-linear weighting function used in the metric, which emphasizes high ozone hours (e.g., 

21 periods in which ozone is greater than ~60 ppb). As an example, consider a sample site in the 

22 intermountain western U.S. region with very high modeled estimates of U.S background (e.g., 

23 seasonal mean of 45 ppb with some days as high as 65 ppb). Even at this high background 

24 location, when the W126 calculation is made for the USB simulation, the resultant annual W126

25 (USB) values are quite low, on the order of 3 ppm-hrs. Sites in the domain with lower U.S. 

26 background levels have even smaller USB W126 values, on the order of the 1 ppm-hrs, which is 

27 consistent with values mentioned in past reviews (USEPA, 2007). Using the counterfactual 

28 scenarios, background ozone has a relatively small impact on W126 levels across the U.S.

29 However, because of the non-linear weighting function used in the W126 calculation, the 

30 sum of the W126 from the USB scenario and the W126 resulting from US anthropogenic sources 

31 will not equal the total W126. In most cases, the sum of those two components will be 

32 substantially less than total W126. As a result, EPA believes it is more informative to estimate 

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
2-23

1 the fractional influence of background ozone to W126 levels. Using a methodology that is 

2 described in more detail in Appendix A, EPA considered the fractional influence of background 

3 ozone on annual W126 levels in four locations. The fractional influence methodology utilizes 

4 the 2007 zero-out modeling but places higher weights on background fractions on days that are 

5 going to contribute most substantially to the yearly W126 value. Figure 2-15 shows the results. 

6 Based on the fractional influence methodology, natural background sources are estimated to 

7 contribute 29-50% of the total modeled W126 with the highest relative influence in the 

8 intermountain western U.S. (e.g., Farmington NM) and the lowest relative influence in the 

9 eastern U.S. (e.g., Atlanta). U.S. background is estimated to contribute 37-65% of the total 

10 modeled W126. The proportional impacts of background are slightly less for the W126 metric 

11 than for seasonal mean MDA8 (discussed in section 2.4.2), because of the sigmoidal weighting 

12 function that places more emphasis on higher ozone days when background fractions are 

13 generally lower.

14 The key conclusion from this cursory analysis is that background ozone may comprise a 

15 non-negligible portion of current W126 levels across the U.S. These fractional influences are 

16 greatest in the intermountain western U.S. and are slightly smaller than the seasonal mean 

17 MDA8 metric. In the counterfactual cases, when non-background sources are completely 

18 removed, the remaining W126 levels are low (< 3 ppm-hrs).

19

20 Figure 2-15. Fractional influence of background sources to W126 levels in four sample 

21 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
2-24

1 2.4.6 Estimated Magnitude of Individual Components of Background O3

2 While local and regional controls of manmade O3 precursor emissions are still expected 

3 to be the most effective mechanism for reducing local O3 levels, an understanding of the relative 

4 contributions of various background elements can be instructive in determining ways to mitigate 

5 the impact of background. This section will utilize the supplemental 2007 air quality modeling 

6 estimates to consider the relative importance of specific elements of background O3. Several 

7 background elements were isolated in either the zero-out or source apportionment modeling. 

8 Appendix A provides more detail on these analyses. In conjunction with the previous analyses 

9 summarized in the ISA, some broad characterizations of the individual components of 

10 background O3 can be developed. 

11 The recent 2007 EPA modeling confirms the importance of methane emissions in 

12 contributing to background O3. Methane has an atmospheric lifetime of about a decade and can 

13 be an important contributor to ozone on longer time scales. Anthropogenic methane emission 

14 sources include agriculture, coal mines, landfills, and natural gas and oil systems. The difference 

15 between the NAB and NB zero-out scenarios provides an estimate of contributions from 

16 international anthropogenic emissions and anthropogenic methane, which is modeled by 

17 reducing model concentrations from present-day values to pre-industrial levels. The ISA (US 

18 EPA, 2013, section 3.4) estimated that roughly half of the difference between the NB and NAB 

19 scenarios resulted from the removal of anthropogenic methane emissions and that the other half 

20 resulted from international anthropogenic emissions of shorter-lived O3 precursors (e.g., NOx 

21 and nmVOC). Figure 2-16 shows the difference in seasonal mean MDA8 O3 levels between the 

22 NB and NAB scenarios. North American seasonal mean background is 6-15 ppb higher than 

23 comparable natural background levels. The most frequent increment is an 8-10 ppb increase 

24 when the methane is increased and the non-North American emissions are re-added. It is not 

25 possible via the EPA 2007 modeling to parse out what fraction of this change is due to emissions 

26 outside of North America, as opposed to the fraction due to anthropogenic methane emissions, 

27 but the modeling suggests that both of these terms have the potential to contribute in an 

28 important way to average background levels in the U.S. 

29 The difference between the NAB and USB scenarios is easier to interpret as it only 

30 involves one change, the inclusion of anthropogenic emissions from the in-domain portion of 

31 Canada and Mexico. These emissions (not shown here, but depicted in Appendix A) contribute 

32 less than 2 ppb to the seasonal mean MDA8 O3 levels over most of the U.S. There are 70 sites, 

33 near an international border, where the modeling estimates Canadian/Mexican seasonal average 

34 impacts of 2-4 ppb. Peak single-day MDA8 impacts from these specific international emissions 

35 sources can approach 25 ppb (e.g., San Diego, Buffalo NY). 

36
2-25

1

2 Figure 2-16. Differences in seasonal mean O3 between the NAB and NB scenarios.

3 Eleven separate source categories were tracked in the source apportionment modeling, 

4 including five boundary condition terms (East, South, West, North, and top) and six emissions 

5 sectors within the domain. The contributions of each of these terms is provided in the Appendix 

6 and summarized below. At most locations, the five model boundary terms contributed an 

7 aggregate 40-60 percent of the total seasonal mean MDA8 O3 across the U.S. The highest 

8 proportional impacts from the boundary conditions are along the coastlines and the 

9 intermountain West. The O3 entering the model domain via the boundary conditions can have a 

10 variety of origins including: a) natural sources of O3 and precursors emanating from outside the 

11 domain, b) anthropogenic sources of O3 precursors emanating from outside the domain, and c) 

12 some fraction of U.S. emissions (natural and anthropogenic) which exit the model domain but 

13 get re-imported into the domain via synoptic-scale recirculation. Accordingly, it is not possible 

14 to relate the boundary condition contribution to any specific background element. The single 

15 largest sector that was tracked in the source apportionment modeling was U.S. anthropogenic 

16 emissions. Figure 2-17 shows the contributions from this sector to seasonal mean MDA8 O3

17 levels. Over most of the U.S. this term contributes 40-60 percent to the total seasonal mean O3. 

18 As discussed in section 2.4.3, these contributions are even higher when only high O3 days are 

19 considered. International shipping emissions, as well as fires and other biogenic emissions also 

20 contribute in a non-negligible way to background O3 over the U.S. The key finding from this 

21 analysis is that air quality planning efforts to reduce anthropogenic methane emissions and 

22 international NOx/nmVOC emissions (e.g., migrating from Asia, Canada, and Mexico; and from 
2-26

1 commercial shipping) have the potential to lower background levels and ease eventual attainment 

2 of the NAAQS. 

3

4

5 Figure 2-17. Percent contribution of U.S. anthropogenic emissions to total seasonal mean 

6 MDA8 O3 levels, based on 2007 source apportionment modeling.

7

8 2.4.7 Summary 

9 For a variety of reasons, it is challenging to present a comprehensive summary of all the 

10 components and implications of background O3. In many forums the term “background” is used 

11 generically and the lack of specificity can lead to confusion as to what sources are being 

12 considered. Additionally, it is well established that the impacts of background sources can vary 

13 greatly over space and time which makes it difficult to present a simple summary of background 

14 O3 levels. Further, background O3 can be generated by a variety of processes, each of which can 

15 lead to differential patterns in space and time, and which often have different regulatory 

16 ramifications. Finally, background O3 is difficult to measure and thus, typically requires air 

17 quality modeling which has inherent uncertainties and potential errors and biases. 

18 That said, EPA believes the following concise and four-stage summary of the 

19 implications of background O3 on the NAAQS review is appropriate, as based on previous 

20 modeling exercises and the more recent EPA analyses summarized herein. First, background O3

21 exists and can comprise a considerable fraction of total seasonal mean O3 and W126 across the 
2-27

1 U.S. Air quality models can estimate the fractional contribution of backgroud sources to total O3

2 in an individual area. The largest absolute values of background (NB, NAB, USB, or 

3 apportionment-based USB) are modeled to occur at locations in the intermountain western U.S. 

4 and are maximized in the spring and early summer seasons. Second, the modeling indicates that 

5 U.S. anthropogenic emission sources are the dominant contributor to the majority of modeled O3

6 exceedances of the NAAQS. Higher O3 days generally have smaller fractional contributions 

7 from background. This finding indicates that the relative importance of background O3 would 

8 increase were O3 concentrations to decrease with a lower level of the O3 NAAQS. Third, while 

9 the majority of modeled O3 exceedances have local and regional emissions as their primary 

10 cause, there can be events where O3 levels approach or exceed 60-75 ppb due to background 

11 sources. These events are relatively infrequent and EPA has policies that could allow for the 

12 exclusion of air quality monitoring data affected by these types of events from design value 

13 calculations. Fourth and finally, the Clean Air Act requires the NAAQS to be set at a level 

14 requisite to protect public health and welfare. Proximity to background levels could be an 

15 additional consideration, but only where it would support a decision based on the health evidence 

16 and analyses. 

17 2.5 REFERENCES 

18 Carlton, A.G.; Pinder, R.W.; Bhave, P.V.; Pouliot, G.A. (2010). To What Extent Can biogenic SOA be Controlled? 

19 Environmental Science & Technology, 44, 3376-3380. 

20 Dunker, A.M.; Yarwood, G; Ortmann, J.P.; Wilson, G.M. (2002). Comparison of source apportionment and source 

21 sensitivity of ozone in a three-dimensional air quality model. Environmental Science & Technology 36: 

22 2953-2964. 

23 Emery, C; Jung, J; Downey, N; Johnson, J; Jimenez, M; Yarwood, G; Morris, R. (2012). Regional and global 

24 modeling estimates of policy relevant background ozone over the United States. Atmos Environ 47: 206-

25 217. http://dx.doi.org/10.1016/j.atmosenv.2011.11.012. 

26 Lefohn, A. S.; Laurence, J. A.; Kohut, R. J. (1988). A comparison of indices that describe the relationship between 

27 exposure to ozone and reduction in the yield of agricultural crops. Atmos. Environ. 22: 1229-1240. 

28 Henderson, B.H., Possiel, N., Akhtar, F., Simon, H.A. (2012). Regional and Seasonal Analysis of North American 

29 Background Ozone Estimates from Two Studies. Available on the Internet at: 

30 http://www.epa.gov/ttn/naaqs/standards/ozone/s_o3_td.html

31 Kemball-Cook, S.; Parrish, D.; Ryerson, T; Nopmongcol, U.; Johnson, J.; Tai, E.; Yarwood, G. (2009). 

32 Contributions of regional transport and local sources to ozone exceedances in Houston and Dallas: 

33 Comparison of results from a photochemical grid model to aircraft and surface measurements. Journal of 

34 Geophysical Research-Atmospheres, 114: D00F02. DOI: 10.1029/2008JD010248. 

35 U.S. Environmental Protection Agency (2005). Technical Support Document for the Final Clean Air Interstate Rule 

36 Air Quality Modeling. Office of Air Quality Planning and Standards, Research Triangle Park, NC, 285pp. 

37 http://www.epa.gov/cair/technical.html. 
2-28

1 U.S. Environmental Protection Agency. (2007). Review of the National Ambient Air Quality Standards for Ozone: 

2 Policy Assessment of Scientific and Technical Information - OAQPS Staff Paper, U.S. Environmental 

3 Protection Agency, Research Triangle Park, NC. EPA 452/R-07-007.

4 U.S. Environmental Protection Agency (2011). Air Quality Modeling Final Rule Technical Support Document. 

5 Office of Air Quality Planning and Standards, Research Triangle Park, NC, 363pp. 

6 http://www.epa.gov/airtransport/CSAPR/techinfo.html. 

7 U.S. Environmental Protection Agency. (2012). Policy Assessment for the Review of the Ozone National Ambient 

8 Air Quality Standards, First External Review Draft, U.S. Environmental Protection Agency, Research 

9 Triangle Park, NC. EPA 452/P-12-002.

10 U.S. Environmental Protection Agency. (2013). Integrated Science Assessment for Ozone and Related 

11 Photochemical Oxidants, U.S. Environmental Protection Agency, Research Triangle Park, NC. 

12 EPA/600/R-10/076. 

13 U.S. Environmental Protection Agency. (2014). Health Risk and Exposure Assessment for Ozone, Second External 

14 Review Draft, U.S. Environmental Protection Agency, Research Triangle Park, NC. EPA-452/P-14-004a. 

15 Wang, HQ; Jacob, DJ; Le Sager, P; Streets, DG; Park, RJ; Gilliland, AB; van Donkelaar, A. (2009). Surface ozone 

16 background in the United States: Canadian and Mexican pollution influences. Atmos Environ 43: 1310-

17 1319. http://dx.doi.org/10.1016/j.atmosenv.2008.11.036. 

18 Zhang, L; Jacob, DJ; Downey, NV; Wood, DA; Blewitt, D; Carouge, CC; Van donkelaar, A; Jones, DBA; Murray, 

19 LT; Wang, Y. (2011). Improved estimate of the policy-relevant background ozone in the United States 

20 using the GEOS-Chem global model with 1/2° × 2/3° horizontal resolution over North America. Atmos 

21 Environ 45: 6769-6776. http://dx.doi.org/10.1016/j.atmosenv.2011.07.054. 
3-1

1 3 Adequacy of the Current Primary Standard 

2 This chapter presents staff’s considerations and conclusions regarding the adequacy of 

3 the current primary O3 NAAQS. In doing so, we pose the following overarching question: 

4

5

6

7 In addressing this overarching question, we pose a series of more specific questions, as discussed 

8 in sections 3.1 through 3.4 below. Section 3.1 presents our consideration of the available 

9 scientific evidence (i.e., evidence-based considerations) about the health effects associated with 

10 short- and long-term O3 exposures. Section 3.2 presents our consideration of available estimates 

11 of O3 exposures and health risks (exposure- and risk-based considerations). Section 3.3 discusses 

12 the advice and recommendations that we have received from the CASAC on the first draft O3

13 PA, and on documents from previous reviews of the O3 NAAQS. Section 3.4 revisits the 

14 overarching question of this section, and presents staff’s conclusions regarding the adequacy of 

15 the current primary O3 NAAQS. 

16 3.1 EVIDENCE-BASED CONSIDERATIONS 

17 This section presents our consideration of the available scientific evidence with regard to 

18 the adequacy of the current O3 standard. Our approach, as summarized in section 1.3.1 above, is 

19 based on the full body of evidence in this review. We use information from the full evidence 

20 base to characterize our confidence in the extent to which O3-attributable effects occur, and the 

21 extent to which such effects are adverse, over the ranges of O3 exposure concentrations evaluated 

22 in controlled human exposure studies and over the distributions of ambient O3 concentrations in 

23 locations where epidemiologic studies have been conducted. In doing so, we recognize that the 

24 available health effects evidence reflects a continuum from relatively high O3 concentrations, at 

25 which scientists generally agree that adverse health effects are likely to occur, through lower 

26 concentrations, at which the likelihood and magnitude of a response become increasingly 

27 uncertain. 

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

1 Section 3.1.1 summarizes a mode of action framework for understanding the effects of 

2 both short- and long-term O3 exposures, based on Chapter 5 of the ISA (U.S. EPA, 2013). 

3 Section 3.1.2 presents our consideration of the evidence for health effects attributable to short4 term and long-term O3 exposures. Section 3.1.3 discusses the adversity of the effects. Section 

5 3.1.4 presents our consideration of evidence with regard to concentrations associated with health 

6 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.17 

8 3.1.1 Modes of Action 

9 Our consideration of the evidence of effects attributable to short-and long-term exposures 

10 and the factors that increase risk in populations and lifestages builds upon evidence about the 

11 modes of action by which O3 exerts effects (U.S. EPA, 2013; section 5.3). Mode of action refers 

12 to a sequence of key events and processes that result in a given toxic effect; elucidation of 

13 mechanisms provides a more detailed understanding of these key events and processes. The 

14 purpose of this section is to describe the key events and pathways that contribute to health effects 

15 resulting from both short-term and long-term exposures to O3. The extensive research carried out 

16 over several decades in humans and animals has yielded numerous studies on mechanisms by 

17 which O3 exerts its effects. It is well-understood that secondary oxidation products, which form 

18 as a result of O3 exposure, initiate numerous responses at the cellular, tissue and whole organ 

19 level of the respiratory system. These responses include the activation of neural reflexes, 

20 initiation of inflammation, alteration of barrier epithelial function, sensitization of bronchial 

21 smooth muscle, modification of lung host defenses, and airways remodeling, as discussed below. 

22 These key events have the potential to affect other organ systems such as the cardiovascular 

23 system. It has been proposed that secondary oxidation products, which are bioactive and 

24 cytotoxic in the respiratory system, are also responsible for systemic effects. Recent studies in 

25 animal models show that inhalation of O3 results in systematic oxidative stress. 

26 Figure 3.1 below, adapted from Figure 5-8 of the ISA (ISA, Section 5.3.10, U.S. EPA, 

27 2013), shows key events in the toxicity pathway of O3 that are described in more detail in 

28 Appendix 3-A. The initial key event in the toxicity pathway of O3 is the formation of secondary 

 

1The term “at-risk populations” includes the factor lifestages, specifically childhood and older adulthood, that are 

experienced by most people over the course of a lifetime, unlike other factors associated with at-risk populations. 
3-3

1 oxidation products in the respiratory tract (ISA, section 5-3, U.S. EPA, 2013). This mainly 

2 involves direct reactions with components of the extracellular lining fluid (ELF). Although the 

3 ELF has inherent capacity to quench (based on individual antioxidant capacity), this attenuative 

capacity can be overwhelmed, especially with exposure to elevated concentrations of O3

2

4 . 

5 The resulting secondary oxidation products transmit signals to the epithelium, pain receptive 

6 nerve fibers and, if present, immune cells (i.e., eosinophils, dendritic cells and mast cells) 

7 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

8 is an implicit 

9 part of this initial key event. 

10

 

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 

superoxides, and a biological system's ability to readily detoxify the reactive intermediates or to repair the resulting 

damage. 

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 
3-4

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

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

3 Figure 3-1 illustrates pathways identified in the ISA by which the array of key events 

4 identified may lead to health effects associated with inhalation exposure to O3. For example, the 

5 activation of neural reflexes, which may be triggered by secondary oxidative products, lead to 

6 lung function decrements and may play a role in extrapulmonary effects such as slow resting 

7 heart rate or bradycardia. Secondary oxidation products have also been implicated in the 

8 initiation of inflammation and inflammation further contributes to O3-mediated oxidative stress. 

9 Alteration of epithelial barrier function may play a role in allergic sensitization and in enhanced 

10 sensitization of bronchial smooth muscle, resulting in airways hyperresponsiveness; genetic 

11 susceptibility has been associated with this pathway. In addition to genetic factors, pre-existing 

12 conditions and diseases, nutritional status, lifestage and co-exposures may affect multiple key 

13 events in Figure 3-1 and contribute to altered risk of O3-induced effects (U.S. EPA, 2013, section 

14 5.4). Evidence also indicates that exposure to O3 modifies innate and adaptive immunity; such 

15 effects can result in both short- and longer-term consequences related to the exacerbation and/or 

16 induction of asthma and to alterations in host defense. Another event, airways remodeling, has 

17 been demonstrated following chronic and/or intermittent exposure to O3 in animal models. 

18 Additionally, there is evidence that O3 exposure results in systemic inflammation and vascular 

19 oxidative/nitrosative stress, which may lead to downstream systemic responses (U.S. EPA, 2013, 

20 section 2.4). 

21 Experimental evidence for such O3-induced changes contributes to our understanding of 

22 the biological plausibility of adverse O3-related health effects, including a range of respiratory 

23 effects as well as effects outside the respiratory system (e.g., cardiovascular effects) (U.S. EPA, 

24 2013, Chapters 6 and 7). The range of respiratory effects that could be mediated by the 

25 secondary oxidation products formed following reactions with O3 include decrements in 

26 pulmonary function; pulmonary inflammation and injury; increased airway permeability; airway 

27 hyperresponsiveness; decreased lung host defense, exacerbation and/or induction of asthma; and 

28 alterations in pulmonary structure and/or development (Figure 3-1, above). These effects are 

29 logically linked to the types of adverse O3-attributable effects evaluated and observed in 

30 epidemiologic studies, including respiratory symptoms, respiratory hospital admissions and 

31 emergency department visits, and premature mortality (U.S. EPA, 2013, Chapters 6 and 7). 

32 Moreover, it has been proposed that some of these key events, including O3-mediated systemic 

33 oxidative stress and activation of neural reflexes, are linked to the extrapulmonary effects of O3

34 that have been noted for decades (U.S. EPA, 2013, section 5.3.2 and 5.3.8). Further, 

35 interindividual variability in the various key events (e.g., due to genetic variants or diet affecting 
3-5

1 antioxidant defenses) illustrated in Figure 3-1 may contribute to differences in susceptibility to 

2 O3 health effects (ISA, section 5.4).

3 3.1.2 Nature of Effects 

4 To what extent does the currently available scientific evidence alter or strengthen 

5 our conclusions from the last review regarding health effects attributable to O3

6 exposure in ambient air? Are previously identified uncertainties reduced or do 

7 important uncertainties remain? 

8 The health effects of ozone are described in detail in the assessment of the evidence 

9 available in this review which is largely consistent with conclusions of past CDs. In some 

10 categories of health effects, there is newly available evidence regarding some aspects of the 

11 effects described in the last review or that strengthens our conclusions regarding aspects of O3

12 toxicity on a particular physiological system (U.S. EPA, 2013, Table 1-1). A sizeable number of 

13 studies on O3 health effects are newly available in this review and are critically assessed in the 

14 ISA as part of the full body of evidence. Based on this assessment, the ISA determined that a 

causal relationship4

exists between short-term exposure to O3 in ambient air515 and effects on the 

respiratory system and that a likely to be causal relationship6

16 exists between long-term exposure 

17 to O3 in ambient air and respiratory effects (U.S. EPA 2013, pp. 1-6 to 1-7). As stated in the 

18 ISA, “[c]ollectively, a very large amount of evidence spanning several decades supports a 

19 relationship between exposure to O3 and a broad range of respiratory effects” (ISA, p. 1-6). 

20 Additionally, the ISA determined likely to be causal relationships exist between short-term 

21 exposures to O3 in ambient air and both total mortality and cardiovascular effects, based on 

22 expanded evidence bases in the current review (U.S. EPA, 2013, pp. 1-7 to 1-8). In the ISA, EPA 

23 additionally determined that the currently available evidence for additional endpoints is 

24 suggestive of causal relationships between short-term (central nervous system effects) and long25 term exposure (cardiovascular effects, central nervous system effects and total mortality) to 

 

4

 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). 

5

 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). 

6

 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). 
3-6

1 ambient O3. Consistent with emphasis in past reviews on O3 health effects for which the evidence 

2 is strongest, we place the greatest emphasis on studies of health effects that have been judged in 

3 the ISA to be caused by, or likely to be caused by, O3 exposures (U.S. EPA, 2013, section 2.5.2). 

4 This section presents our consideration of the evidence for health effects attributable to 

5 O3 exposures, including respiratory morbidity and mortality effects attributable to short- and 

6 long-term exposures, and cardiovascular system effects (including mortality) and total mortality 

7 attributable to short-term exposures. We focus particularly on considering the extent to which the 

8 scientific evidence available in the current review has been strengthened since the last review, 

9 and the extent to which important uncertainties and limitations in the evidence from the last 

10 review have been addressed. In section 3.1.2.2, we then consider the extent to which the 

11 available evidence indicates health effects may be attributable to ambient O3 concentrations 

12 likely to be allowed by the current O3 NAAQS. In this section, we address the following specific 

13 question for each category of health effects considering the evidence available in the 2008 

14 review of the standard as well as evidence that has become available since then. The ISA 

15 summarizes the longstanding body of evidence for O3 respiratory effects as follows (U.S. EPA, 

16 2013, p. 1-5). 

17 The clearest evidence for health effects associated with exposure to O3 is provided 

18 by studies of respiratory effects. Collectively, a very large amount of evidence 

19 spanning several decades supports a relationship between exposure to O3 and a 

20 broad range of respiratory effects (see Section 6.2.9 and Section 7.2.8). The 

21 majority of this evidence is derived from studies investigating short-term 

22 exposures (i.e., hours to weeks) to O3, although animal toxicological studies and 

23 recent epidemiologic evidence demonstrate that long-term exposure (i.e., months 

24 to years) may also harm the respiratory system. 

25 The extensive body of evidence supporting a causal relationship between short-term O3

26 exposures and respiratory effects is discussed in detail in Chapter 6 of the ISA (U.S. EPA, 2013), 

27 while evidence for respiratory effects associated with long-term or repeated O3 exposures are 

28 discussed in chapter 7 of that document (U.S., EPA, 2013). 

29 3.1.2.1 Respiratory Effects – Short-term Exposures 

30 To what extent does the currently available scientific evidence, including related 

31 uncertainties, strengthen or alter our understanding from the last review of respiratory 

32 effects attributable to short-term O3 exposures? 
3-7

1 The 2006 O3 AQCD concluded that there was clear, consistent evidence of a causal 

2 relationship between short-term O3 exposure and respiratory effects (U.S. EPA, 2006b). This 

3 conclusion was substantiated by evidence from controlled human exposure and toxicological 

4 studies indicating a range of respiratory effects in response to short-term O3 exposures, including 

5 pulmonary function decrements and increases in respiratory symptoms, lung inflammation, lung 

6 permeability, and airway hyperresponsiveness. Toxicological studies provided additional 

7 evidence for O3-induced impairment of host defenses. Combined, these findings from 

8 experimental studies provided support for epidemiologic evidence, in which short-term increases 

9 in ambient O3 concentration were consistently associated with decreases in lung function in 

10 populations with increased outdoor exposures, especially children with asthma and healthy 

11 children; increases in respiratory symptoms and asthma medication use in children with asthma; 

12 and increases in respiratory-related hospital admissions and asthma-related ED visits (U.S. EPA, 

13 2013, pp. 6-1 to 6-2). 

14 As discussed in detail in the ISA (U.S. EPA, 2013, section 6.2.9), studies evaluated since 

15 the completion of the 2006 O3 AQCD support and expand upon the strong body of evidence that, 

16 in the last review, indicated a causal relationship between short-term O3 exposures and 

17 respiratory health effects. Recent controlled human exposure studies conducted in young, healthy 

18 adults with moderate exertion have reported FEV1 decrements and pulmonary inflammation 

19 following prolonged exposures to O3 concentrations as low as 60 ppb, and respiratory symptoms 

20 following exposures to concentrations as low as 70 ppb. Epidemiologic studies provide evidence 

21 that increases in ambient O3 exposures can result in lung function decrements, increases in 

22 respiratory symptoms, and pulmonary inflammation in children with asthma; increases in 

23 respiratory-related hospital admissions and emergency department visits; and increases in 

24 respiratory mortality. Some of these studies report such associations even for O3 concentrations 

25 at the low end of the distribution of daily concentrations. Recent epidemiologic studies report 

26 that associations with respiratory morbidity and mortality are stronger during the warm/summer 

27 months and remain robust after adjustment for copollutants. Recent toxicological studies 

28 reporting O3-induced inflammation, airway hyperresponsiveness, and impaired lung host defense 

29 continue to support the biological plausibility and modes of action for the O3-induced respiratory 

30 effects observed in the controlled human exposure and epidemiologic studies. Further support is 

31 provided by recent studies that found O3-associated increases in indicators of airway 
3-8

1 inflammation and oxidative stress in children with asthma (U.S. EPA, 2013, section 6.2.9). 

2 Together, epidemiologic and experimental studies support a continuum of respiratory effects 

3 associated with O3 exposure that can result in respiratory-related emergency department visits, 

4 hospital admissions, and/or mortality (U.S. EPA, 2013, section 6.2.9). 

5 Across respiratory endpoints, evidence indicates antioxidant capacity may modify the 

6 risk of respiratory morbidity associated with O3 exposure (U.S. EPA, 2013, section 6.2.9, p. 6-

7 161). The potentially elevated risk of populations with diminished antioxidant capacity and the 

8 reduced risk of populations with sufficient antioxidant capacity identified in epidemiologic 

9 studies are strongly supported by similar findings from controlled human exposure studies and 

10 by evidence that characterizes O3-induced decreases in intracellular antioxidant levels as a mode 

11 of action for downstream effects. 

12 We describe key aspects of this evidence below with regard to lung function decrements; 

13 pulmonary inflammation, injury, and oxidative stress; airway hyperresponsiveness; respiratory 

14 symptoms and medication use; lung host defense; allergic and asthma-related responses; hospital 

15 admissions and emergency department visits; and respiratory mortality. 

16 Lung Function Decrements 

In the 2008 review, a large number of controlled human exposure studies717 reported O3-

18 induced lung function decrements in young, healthy adults engaged in intermittent, moderate 

19 exertion following 6.6 hour exposures to O3 concentrations at or above 80 ppb. Although two 

20 studies also reported effects following exposures to lower concentrations, an important 

21 uncertainty in the last review was the extent to which exposures to O3 concentrations below 80 

22 ppb result in lung function decrements. In addition, in the last review epidemiologic panel 

23 studies had reported O3-associated lung function decrements in a variety of different populations 

24 (e.g., children, outdoor workers) likely to experience increased exposures. In the current review, 

25 additional controlled human exposure studies are available that have evaluated exposures to O3

26 concentrations of 60 or 70 ppb. The available evidence from controlled human exposure and 

27 panel studies is assessed in detail in the ISA (U.S. EPA, section 6.2.1) and is summarized below. 

 

7

 The controlled human exposure studies discussed in this PA utilize only healthy adult subjects. In the 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, section 6.2.4 and 6.5). Thus, the conclusions about the occurrence of lung function decrements 

that follow generally apply to children as well as adults. 
3-9

1 Controlled exposures to O3 concentrations that can be found in the ambient air can result 

2 in a number of lung function effects, including decreased inspiratory capacity; mild 

3 bronchoconstriction; and rapid, shallow breathing patterns during exercise. Reflex inhibition of 

4 inspiration results in a decrease in forced vital capacity (FVC) and total lung capacity (TLC) and, 

5 in combination with mild bronchoconstriction, contributes to a decrease in the forced expiratory 

6 volume in 1 second (FEV1) (U.S. EPA, 2013, section 6.2.1.1). Accumulating evidence indicates 

7 that such effects are mediated by activation of sensory nerves, resulting in the involuntary 

8 truncation of inspiration and a mild increase in airway obstruction due to bronchoconstriction 

9 (U.S. EPA, 2013, section 5.3.10). 

10 Data from controlled human exposure studies indicate that increasing the duration of O3

11 exposures and increasing ventilation rates decreases the O3 exposure concentrations required to 

12 impair lung function. Ozone exposure concentrations well above those typically found in 

13 ambient air are required to impair lung function in healthy resting adults, while exposure to O3

14 concentrations at or below those in the ambient air have been reported to impair lung function in 

15 healthy adults exposed for longer durations while undergoing intermittent, moderate exertion 

16 (U.S. EPA, 2012a, section 6.2.1.1). With repeated O3 exposures over several days, FEV1

17 responses become attenuated in both healthy adults and adults with mild asthma, though this 

18 attenuation of response is lost after about a week without exposure (U.S. EPA, 2013, section 

19 6.2.1.1; page 6-27). 

20 When considering controlled human exposures studies of O3-induced lung function 

21 decrements we evaluate both group mean changes in lung function and the interindividual 

22 variability in the magnitude of responses. To the extent studies report statistically significant 

23 decrements in mean lung function following O3 exposures after controlling for other factors, we 

24 have more confidence that measured decrements are due to the O3 exposure itself, rather than 

25 chance alone. As discussed below, group mean changes in lung function are often small, 

26 especially following exposures to relatively low O3 concentrations (e.g., 60 ppb). However, even 

27 when group mean decrements in lung function are small, some individuals could experience 

28 decrements that are “clinically meaningful” (Pellegrino et al., 2005; ATS, 1991; EPA, 2010) 

29 with respect to criteria for spirometric testing, and/or that could be considered “adverse” with 

30 respect to public health policy decisions. See section 3.1.3 below. 
3-10

1 At the time of the last review, a number of controlled human exposure studies had 

2 reported lung function decrements in young, healthy adults following prolonged (6.6-hour) 

3 exposures while at moderate exertion to O3 concentrations at and above 80 ppb. In addition, 

4 there were two controlled human exposure studies by Adams (2002, 2006) that examined lung 

5 function effects following exposures to O3 concentrations of 60 ppb. The EPA’s analysis of the 

6 data from the Adams (2006) study reported a small but statistically significant O3-induced 

7 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, 2007). 8

8 Further 

9 examination of the post-exposure FEV1 data, and mean data for other time points and other 

10 concentrations, indicated that the temporal pattern of the response to 60 ppb O3 was generally 

11 consistent with the temporal patterns of responses to higher O3 concentrations in this and other 

12 studies. (75 FR 2950, January 19, 2010). This suggested a pattern of response following 

13 exposures to 60 ppb O3 that was consistent with a dose-response relationship, rather than random 

14 variability. See also State of Mississippi v. EPA, 723 F. 3d at 259 (upholding EPA’s 

15 interpretation of the Adams studies). 

16 Figure 6-1 in the ISA summarizes the currently available evidence from multiple 

17 controlled human exposure studies evaluating group mean changes in FEV1 following prolonged 

18 O3 exposures (i.e., 6.6 hours) in young, healthy adults engaged in moderate levels of physical 

19 activity (U.S. EPA, 2012, section 6.2.1.1). With regard to the group mean changes reported in 

20 these studies, the ISA specifically notes the following (U.S. EPA, 2012a, section 6.2.1.1, Figure 

21 6-1): 

22 1. Prolonged exposure to 40 ppb O3 results in a small decrease in group mean FEV1 that is 

23 not statistically different from responses following exposure to filtered air (Adams, 2002; 

24 Adams, 2006). 

25 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;9

26 Schelegle et al., 

27 2009; Kim et al., 2011). Based on data from multiple studies, the weighted average group 

28 mean decrement was 2.7%. In some analyses, these group mean decrements in lung 

 

8

Adams (2006a) did not find effects on FEV1 at 60 ppb to be statistically significant. In an analysis of the Adams 

(2006a) 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. 

9

 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 ( ) = 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 ( ) of 23 L/min per m2 BSA relative to other studies with which it is 

listed having the target ( ) of 20 L/min per m2 BSA. The facemask exposure is not expected to affect the FEV1

responses relative to a chamber exposure. 
3-11

1 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).10 2

3 3. Prolonged exposure to an average O3 concentration of 70 ppb results in a statistically 

4 significant group mean decrement in FEV1 of about 6% (Schelegle et al., 2009). 

5 4. Prolonged square-wave exposure to average O3 concentrations of 80 ppb, 100 ppb, or 120 

6 ppb O3 results in statistically significant group mean decrements in FEV1 ranging from 6 

7 to 8%, 8 to 14%, and 13 to 16%, respectively (Folinsbee et al., 1988; Horstman et al., 

8 1990; McDonnell et al., 1991; Adams, 2002; Adams, 2003; Adams, 2006). 

9 As illustrated in Figure 6-1 of the ISA, there is a smooth dose-response curve without 

10 evidence of a threshold for exposures between 40 and 120 ppb O3 (U.S. EPA, 2012a, Figure 6-

11 1). When these data are taken together, the ISA concludes that “mean FEV1 is clearly decreased 

12 by 6.6-h exposures to 60 ppb O3 and higher concentrations in [healthy, young adult] subjects 

13 performing moderate exercise” (U.S. EPA, 2012a, p. 6-9). 

14 With respect to interindividual variability in lung function, in an individual with 

15 relatively “normal” lung function, with recognition of the technical and biological variability in 

16 measurements, within-day changes in FEV1 of ≥ 5% are clinically meaningful (Pellegrino et al., 

17 2005; ATS, 1991). The ISA (U.S. EPA, 2013, section 6.1.) focuses on individuals with >10% 

18 decrements in FEV1 for two reasons. A 10% FEV1 decrement is accepted by the American 

19 Thoracic Society (ATS) as an abnormal response and a reasonable criterion for assessing 

20 exercise-induced bronchoconstriction (Dryden et al., 2010; ATS, 2000a). (U.S. EPA, 2013, 

21 section 6.2.1.1). Also, some individuals in the Schelegle et al. (2009) study experienced 5-10% 

22 FEV1 decrements following exposure to filtered air. 

23 In previous NAAQS reviews, the EPA has made judgments regarding the potential 

implications for individuals experiencing FEV1 decrements of varying degrees of severity.11 24 For 

25 people with lung disease, the EPA judged that moderate functional decrements (e.g., FEV1 

26 decrements > 10 percent but < 20 percent, lasting up to 24 hours) would likely interfere with 

 

10Adams (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. 

11Such 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 2007 Staff Paper (U.S. EPA, 2007). 
3-12

1 normal activity for many individuals, and would likely result in more frequent use of medication 

2 (75 FR 2973, January 19, 2010). In previous reviews CASAC has endorsed these conclusions. In 

3 the context of standard setting, in the last review O3 review CASAC indicated that it is 

4 appropriate to focus on the lower end of the range of moderate functional responses (e.g., FEV1

5 decrements 10 percent) when estimating potentially adverse lung function decrements in 

6 people with lung disease, especially children with asthma (Henderson, 2006c; transcript of 

7 CASAC meeting, day 8/24/06, page 149). More specifically, CASAC stated that “[a] 10% 

8 decrement in FEV1 can lead to respiratory symptoms, especially in individuals with pre-existing 

9 pulmonary or cardiac disease. For example, people with chronic obstructive pulmonary disease 

10 have decreased ventilatory reserve (i.e., decreased baseline FEV1) such that a ≥ 10% decrement 

11 could lead to moderate to severe respiratory symptoms” (Samet, 2011). Therefore, in considering 

12 interindividual variability in O3-induced lung function decrements in the current review, we also 

13 focus on the extent to which individuals were reported to experience FEV1 decrements of 10% or 

14 greater. 

15 New studies (Schelegle et al., 2009; Kim et al., 2011) add to the previously available 

16 evidence for interindividual variability in the responses of healthy adults following exposures to 

17 O3. Following prolonged exposures to 80 ppb O3 while at moderate exertion, the proportion of 

18 healthy adults experiencing FEV1 decrements greater than 10% was 17% by Adams (2006a), 

19 26% by McDonnell (1996), and 29% by Schelegle et al. (2009). Following exposures to 60 ppb 

20 O3, that proportion was 20% by Adams (2002), 3% by Adams (2006a), 16% by Schelegle et al. 

21 (2009), and 5% by Kim et al. (2011). Based on these studies, the weighted average proportion of 

22 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).12 23 The ISA notes that 

24 responses within an individual tend to be reproducible over a period of several months, 

25 indicating that interindividual differences reflect differences in intrinsic responsiveness. Given 

26 this, the ISA concludes that “a considerable fraction” of healthy individuals experience clinically 

 

12The 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, 2012, 

section 6.2.1.1). 
3-13

1 meaningful decrements in lung function when exposed for 6.6 hours to 60 ppb O3 during quasi 

2 continuous, moderate exertion (U.S. EPA, 2013, section 6.2.1.1, p. 6-20).

3 As discussed above (Figure 3-1) and in the ISA (U.S EPA, 2013, Section 5.3.2), 

4 secondary oxidation products formed following O3 exposures can activate neural reflexes leading 

5 to decreased lung function. Two new quantitative models, discussed in section 6.2.1.1 of the ISA 

6 (U.S. EPA, 2013, p. 6-15), make use of the concept of oxidant stress to estimate the occurrence 

7 of lung function decrements following exposures to relatively low O3 concentrations (McDonnell 

8 et al., 2012; Schelegle et al., 2012). These models reflect the protective effect of antioxidants in 

9 the ELF at lower ambient O3 concentrations, and include a threshold related to an integrated dose 

10 rate. 

11 McDonnell et al. (2012) and Schelegle et al. (2012) developed models using data on O3

12 exposure concentrations, ventilation rates, duration of exposures, and lung function responses 

13 from a number of controlled human exposure studies. The McDonnell et al. (2012) and Schelegle 

14 et al. (2012) studies analyzed large datasets to fit compartmental models that included the 

15 concept of a dose of onset in lung function response or a response threshold based upon the 

16 inhaled O3 dose. The first compartment in the McDonnell et al. (2012) model considers the level 

17 of oxidant stress in response to O3 exposure to increase over time as a function of dose rate 

18 (C× ) and decrease by clearance or metabolism over time. In the second compartment of the 

19 McDonnell model, once oxidant stress reaches a threshold level the decrement in FEV1 increases 

20 as a sigmoid-shaped function. In the Schelegle et al. (2012) model, a first compartment acts as a 

21 reservoir in which oxidant stress builds up until the dose of onset, at which time it spills over into 

22 a second compartment. The second compartment is identical to the first compartment in 

23 McDonnell et al. (2012) model. The oxidant levels in the second compartment were multiplied 

24 by a responsiveness coefficient to predict FEV1 responses for the Schelegle et al. (2012) model. 

25 The McDonnell et al. (2012) model was fit to a large dataset consisting of the FEV1

26 responses of 741 young, healthy adults (18-35 years of age) from 23 individual controlled 

27 exposure studies. Concentrations across individual studies ranged from 40 ppb to 400 ppb, 

28 activity level ranged from rest to heavy exercise, duration of exposure was from 2 to 7.6 hours. 

29 The extension of the McDonnell et al. (2012) model to children and older adults is discussed in 

30 section 6.2.4 of U.S. EPA (2014). Schelegle et al. (2012) also analyzed a large dataset with 

31 substantial overlap to that used by McDonnell et al. (2012). The Schelegle et al. (2012) model 

32 was fit to the FEV1 responses of 220 young healthy adults (taken from a dataset of 704 

33 individuals) from 21 individual controlled exposure studies. The resulting empirical models can 

34 estimate the frequency distribution of individual responses for any exposure scenario as well as 
3-14

1 summary measures of the distribution such as the mean or median response and the proportions 

2 of individuals with FEV1 decrements > 10%, 15%, and 20%. 

3 The predictions of the McDonnell and Schelegle models are consistent with the observed 

4 results from the individual studies of O3-induced FEV1 decrements. Specifically, the model 

5 developed by McDonnell et al. (2012) predicts that 9% of healthy exercising adults would 

6 experience FEV1 decrements greater than 10% following 6.6 hour exposure to 60 ppb O3, and 

7 that 22% would experience such decrements following exposure to 80 ppb O3 (U.S. EPA, 2013, 

8 p. 6-18 and Figure 6-3). The model developed by Schelegle et al. (2012) predicts that, for a 

9 prolonged (6.6 hours) O3 exposure with moderate, quasi continuous exercise, the average dose of 

10 onset for FEV1 decrement would be reached following 4 to 5 hours of exposure to 60 ppb, and 

11 following 3 to 4 hours of exposure to 80 ppb. However, 14% of the individuals had a dose of 

12 onset that was less than 40% of the average. Those individuals would reach their dose of onset 

13 following 1 to 2 hours of exposure to 50 to 80 ppb O3 (U.S. EPA, 2013, p. 6-16), which is 

14 consistent with the threshold FEV1 responses reported by McDonnell et al. (2012). 

Epidemiologic studies13 15 have consistently linked short-term increases in ambient O3

16 concentrations with lung function decrements in diverse populations and lifestages, including 

17 children attending summer camps, adults exercising or working outdoors, and groups with pre18 existing respiratory diseases such as asthmatic children (U.S. EPA, 2013, section 6.2.1.2). Some 

19 of these studies reported ozone-associated lung function decrements accompanied by respiratory 

symptoms14 20 in asthmatic children (Just et al., 2002; Mortimer et al., 2002; Ross et al., 2002; 

21 Gielen et al., 1997; Romieu et al., 1997; Thurston et al., 1997; Romieu et al., 1996). In contrast, 

22 studies of children in the general population have reported similar O3-associated lung function 

23 decrements but without accompanying respiratory symptoms (Ward et al., 2002; Gold et al., 

24 1999; Linn et al., 1996) (U.S. EPA, 2013, section 6.2.1.2). 

25 Several panel studies reported that associations with lung function decrements persisted 

26 at relatively low ambient O3 concentrations. For outdoor recreation or exercise, associations were 

27 reported in analyses restricted to 1-hour average O3 concentrations less than 80 ppb (Spektor et 

28 al., 1988a; Spektor et al., 1988b), 60 ppb (Brunekreef et al., 1994; Spektor et al., 1988a), and 

29 50 ppb (Brunekreef et al., 1994). Among outdoor workers, Brauer et al. (1996) found a robust 

30 association using daily 1-hour max O3 concentrations less than 40 ppb. Ulmer et al. (1997) found 

31 a robust association in schoolchildren using 30-minute maximum O3 concentrations less than 

32 60 ppb. For 8-hour average O3 concentrations, associations with lung function decrements in 

 

13 Unless otherwise specified, the epidemiologic studies discussed in this PA evaluate only adults. 

14 Reversible loss of lung function in combination with the presence of symptoms meets the ATS definition of 

adversity (ATS, 2000). 
3-15

1 children with asthma were found to persist at concentrations less than 80 ppb in a U.S. multicity 

2 study (Mortimer et al., 2002) and less than 51 ppb in a study conducted in the Netherlands 

3 (Gielen et al., 1997). 

4 Studies investigating the effects of short-term exposure to O3 provided information on 

5 potential confounding by copollutants such as PM2.5, PM10, NO2, or SO2. These studies varied in 

6 how they evaluated confounding. Some studies of subjects exercising outdoors indicated that 

7 ambient concentrations of copollutants such as NO2, SO2, or acid aerosol were low, and thus not 

8 likely to confound associations observed for O3 (Hoppe et al., 2003; Brunekreef et al., 1994; 

9 Hoek et al., 1993). In other studies of children with increased outdoor exposures, O3 was 

10 consistently associated with decreases in lung function, whereas other pollutants such as PM2.5, 

11 sulfate, and acid aerosol individually showed variable associations across studies (Thurston et 

12 al., 1997; Castillejos et al., 1995; Berry et al., 1991; Avol et al., 1990; Spektor et al., 1988a). 

13 Studies that conducted copollutant modeling generally found O3-associated lung function 

14 decrements to be robust (i.e., most copollutant-adjusted effect estimates fell within the 95% CI of 

15 the single-pollutant effect estimates) (U.S. EPA, 2013, Figure 6-10 and Table 6-14). Most O3

16 effect estimates for lung function were robust to adjustment for temperature, humidity, and 

17 copollutants such as PM2.5, PM10, NO2, or SO2. Although examined in only a few epidemiologic 

18 studies, O3 also remained associated with decreases in lung function with adjustment for pollen 

19 or acid aerosols (U.S. EPA, 2013, section 6.2.1.2). 

20 Several epidemiologic studies demonstrated the protective effects of vitamin E and 

21 vitamin C supplementation, and increased dietary antioxidant intake, on O3-induced lung 

22 function decrements (Romieu et al., 2002) (U.S. EPA, 2013, Figure 6-7 and Table 6-8). These 

23 results provide support for the new, quantitative models (McDonnell et al., 2012; Schelegle et 

24 al., 2012), discussed above, which make use of the concept of oxidant stress to estimate the 

25 occurrence of lung function decrements following exposures to relatively low O3 concentrations. 

26 In conclusion, new information from controlled human exposure studies considerably 

27 strengthens the evidence and reduces the uncertainties, relative to the evidence that was available 

28 at the time of the 2008 review, regarding the presence and magnitude of lung function 

29 decrements in healthy adults following prolonged exposures to O3 concentrations below 80 ppb. 

30 As discussed in Section 6.2.1.1 in the ISA (EPA, 2013, p. 6-12), there is information available 

31 from four separate studies that evaluated exposures to 60 ppb O3 (Kim et al., 2011; Schelegle et 

32 al., 2009; Adams 2002; 2006). Although not consistently statistically significant, group mean 

33 FEV1 decrements following exposures to 60 ppb O3 are consistent among studies. Moreover, as 

34 is illustrated in Figure 6-1 of the ISA (U.S. EPA, 2013), the group mean FEV1 responses at 

35 60 ppb fall on a smooth intake dose-response curve for exposures between 40 and 120 ppb O3. 

36 These studies also indicate that, on average, 10% of young, healthy adults experience clinically 
3-16

1 meaningful decrements in lung function when exposed for 6.6 hours to 60 ppb O3 during 

2 intermittent, moderate exertion. One recent study has also reported statistically significant 

3 decrements following exposures to 70 ppb O3 (Schelegle et al., 2009). Predictions from newly 

4 developed quantitative models, based on the concept that O3-induced oxidation results in lung 

5 function decrements, are consistent with these experimental results. Additionally, as discussed in 

6 more detail in section 3.1.4 below, epidemiologic studies continue to provide evidence of lung 

7 function decrements in people who are active outdoors, including people engaged in outdoor 

8 recreation or exercise, children, and outdoor workers, at low ambient O3 concentrations. While 

9 few new epidemiologic studies of O3-associated lung function decrements are available in this 

10 review, previously available studies have reported associations with decrements, including at 

11 relatively low ambient O3 concentrations. 

12 Pulmonary Inflammation, Injury, and Oxidative Stress 

13 Ozone exposures result in increased respiratory tract inflammation and epithelial 

14 permeability. Inflammation is a host response to injury, and the induction of inflammation is 

15 evidence that injury has occurred. Oxidative stress has been shown to play a key role in initiating 

16 and sustaining O3-induced inflammation. Secondary oxidation products formed as a result of 

17 reactions between O3 and components of the ELF can increase the expression of molecules (i.e., 

18 cytokines, chemokines, and adhesion molecules) that can enhance airway epithelium 

19 permeability (U.S. EPA, 2013, Sections 5.3.3 and 5.3.4). As discussed in detail in the ISA (U.S. 

20 EPA, 2013, section 6.2.3), O3 exposures can initiate an acute inflammatory response throughout 

21 the respiratory tract that has been reported to persist for at least 18-24 hours after exposure. 

22 Inflammation induced by exposure of humans to O3 can have several potential outcomes: 

23 (1) inflammation induced by a single exposure (or several exposures over the course of a 

24 summer) can resolve entirely; (2) continued acute inflammation can evolve into a chronic 

25 inflammatory state; (3) continued inflammation can alter the structure and function of other 

26 pulmonary tissue, leading to diseases such as asthma; (4) inflammation can alter the body’s host 

27 defense response to inhaled microorganisms, particularly in potentially at-risk populations or 

28 lifestages such as the very young and old; and (5) inflammation can alter the lung’s response to 

29 other agents such as allergens or toxins (U.S. EPA, 2013, Section 6.2.3). Thus, lung injury and 

30 the resulting inflammation provide a mechanism by which O3 may cause other more serious 

31 morbidity effects (e.g., asthma exacerbations). 
3-17

1 In the last review, controlled human exposure studies reported O3-induced airway 

2 inflammation following exposures at or above 80 ppb. In the current review, the link between O3

3 exposures and airway inflammation and injury has been evaluated in additional controlled human 

4 exposure studies, as well as in recent epidemiologic studies. Controlled human exposure studies 

5 have generally been conducted in young, healthy adults or in adults with asthma using lavage 

6 (proximal airway and bronchoalveolar), bronchial biopsy, and more recently, induced sputum. 

These studies have evaluated one or more indicators of inflammation, including neutrophil15 7

8 (PMN) influx, markers of eosinophilic inflammation, increased permeability of the respiratory 

9 epithelium, and/or prevalence of proinflammatory molecules (U.S. EPA, 2013, section 6.2.3.1). 

10 Epidemiologic studies have generally evaluated associations between ambient O3 and markers of 

11 inflammation and/or oxidative stress, which plays a key role in initiating and sustaining 

12 inflammation (U.S. EPA, 2013, section 6.2.3.2). 

13 There is an extensive body of evidence from controlled human exposure studies 

14 indicating that short-term exposures to O3 can cause pulmonary inflammation. Previously 

15 available evidence indicated that O3 causes an inflammatory response in the lungs (U.S. EPA, 

16 1996a). A single acute exposure (1-4 hours) of humans to moderate concentrations of O3 (200-

17 600 ppb) while exercising at moderate to heavy intensities resulted in a number of cellular and 

18 biochemical changes in the lung, including inflammation characterized by increased numbers of 

19 PMNs, increased permeability of the epithelial lining of the respiratory tract, cell damage, and 

20 production of proinflammatory molecules (i.e., cytokines and prostaglandins, U.S. EPA, 2006b).

21 A meta-analysis of 21 controlled human exposure studies (Mudway and Kelly, 2004) using 

22 varied experimental protocols (80-600 ppb O3 exposures; 1-6.6 hours exposure duration; light to 

23 heavy exercise; bronchoscopy at 0-24 hours post-O3 exposure) reported that PMN influx in 

24 healthy subjects is linearly associated with total O3 dose. Animal toxicological studies also 

25 provided evidence for increases in inflammation and permeability in rabbits at levels as low as 

26 100 ppb O3 (Section 2.5.3.1, ISA, U.S. EPA, 2013). 

 

15 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. 
3-18

1 Several studies, including one published since the last review (Alexis et al., 2010), have 

2 reported O3-induced increases in PMN influx and permeability following exposures at or above 

3 80 ppb (Alexis et al., 2010; Peden et al., 1997; Devlin et al., 1991), and eosinophilic 

4 inflammation following exposures at or above 160 ppb (Scannell et al., 1996; Peden et al., 1997; 

5 Hiltermann et al., 1999; Vagaggini et al., 2002). In addition, one recent controlled human 

6 exposure study has reported O3-induced PMN influx following exposures of healthy adults to O3

7 concentrations of 60 ppb (Kim et al., 2011), the lowest concentration at which inflammatory 

8 responses have been evaluated in human studies. 

9 As with FEV1 responses to O3, inflammatory responses to O3 are generally reproducible 

10 within individuals, with some individuals experiencing more severe O3-induced airway 

11 inflammation than indicated by group averages (Holz et al., 2005; Holz et al., 1999). Unlike O3-

12 induced decrements in lung function, which are attenuated following repeated exposures over 

13 several days (U.S. EPA, 2013, section 6.2.1.1), some markers of O3-induced inflammation and 

14 tissue damage remain elevated during repeated exposures, indicating ongoing damage to the 

15 respiratory system (U.S. EPA, 2013, section 6.2.3.1). 

16 Most controlled human exposure studies have reported that asthmatics experience larger 

17 O3-induced inflammatory responses than non-asthmatics. Specifically, asthmatics exposed to 

18 200 ppb O3 for 4-6 hours with exercise show significantly more neutrophils in bronchoalveolar 

19 lavage fluid (BALF) than similarly exposed healthy individuals (Scannell et al., 1996; Basha et 

20 al., 1994). Bosson et al. (2003) reported significantly greater expression of a variety of pro21 inflammatory cytokines in asthmatics, compared to healthy subjects, following exposure to 

22 200 ppb O3 for 2 hours. In addition, research available in the last review, combined with a recent 

23 study newly available in this review, indicates that pretreatment of asthmatics with 

24 corticosteroids can prevent the O3-induced inflammatory response in induced sputum, though 

25 pretreatment did not prevent FEV1 decrements (Vagaggini et al., 2001; 2007). In contrast, 

26 Stenfors et al. (2002) did not detect a difference in the O3-induced increases in neutrophil 

27 numbers between 15 subjects with mild asthma and 15 healthy subjects by bronchial wash at the 

28 6 hours postexposure time point, although the neutrophil increase in the asthmatic group was on 

29 top of an elevated baseline. 

30 In people with allergic airway disease, including people with rhinitis and asthma, 

31 evidence available in the last review indicated that proinflammatory mediators also cause 
3-19

1 accumulation of eosinophils in the airways (Bascom et al., 1990; Jorres et al., 1996; Peden et 

2 al.,1995 and 1997; Frampton et al., 1997; Michelson et al., 1999; Hiltermann et al., 1999; Holz et 

3 al., 2002; Vagaggini et al., 2002). The eosinophil, which increases inflammation and allergic 

4 responses, is the cell most frequently associated with exacerbations of asthma (75 CFR 2969, 

5 January 19, 2010). 

6 Studies reporting inflammatory responses and markers of lung injury have clearly 

7 demonstrated that there is important variation in the responses of exposed subjects (75 FR 2953, 

8 January 19, 2010). Some individuals also appear to be intrinsically more susceptible to increased 

9 inflammatory responses (Holz et al., 2005). In healthy adults exposed to each 80 and 100 ppb O3, 

10 Devlin et al. (1991) observed group average increases in neutrophilic inflammation of 2.1- and 

11 3.8-fold, respectively. However, there was a 20-fold range in inflammatory responses between 

12 individuals at both concentrations. Relative to an earlier, similar study conducted at 400 ppb 

13 (Koren et al., 1989), Devlin et al. (1991) noted that although some of the study population 

14 showed little or no increase in inflammatory and cellular injury indicators analyzed after 

15 exposures to lower levels of O3 (i.e., 80 and 100 ppb), others had changes that were as large as 

16 those seen when subjects were exposed to 400 ppb O3. The data suggest that as a whole the 

17 healthy population, on average, may have small inflammatory responses to near-ambient levels 

18 of O3, though there may be a significant subpopulation that is very sensitive to low levels of O3. 

19 Devlin et al. (1991) expressed the view that ‘‘susceptible subpopulations such as the very young, 

20 elderly, and people with pulmonary impairment or disease may be even more affected.’’ 

21 A number of studies report that O3 exposures increase epithelial cell permeability.

22 Increased BALF protein, suggesting O3-induced changes in epithelial permeability, has been 

23 reported at 1 hour and 18 hours postexposure (Devlin et al., 1997; Balmes et al., 1996). A meta24 analysis of results from 21 publications (Mudway and Kelly, 2004a) for varied experimental 

25 protocols (80-600 ppb O3; 1-6.6 hours duration; light to heavy exercise; bronchoscopy at 0-24 

26 hours post-O3 exposure; healthy subjects), showed that increased BALF protein is associated 

27 with total inhaled O3 dose (i.e., the product of O3 concentration, exposure duration, and ). As 

28 noted in the 2009 PM ISA (U.S. EPA, 2009), it has been postulated that changes in permeability 

29 associated with acute inflammation may provide increased access of inhaled antigens, particles, 

30 and other inhaled substances deposited on lung surfaces to the smooth muscle, interstitial cells, 

31 immune cells underlying the epithelium, and the blood (U.S. EPA, 2013, sections 5.3.4, 5.3.5).
3-20

1 Animal toxicology studies have provided some support for this hypothesis (Adamson and 

2 Prieditis, 1995; Chen et al 2006), though these studies did not specifically evaluate O3 exposures 

3 (U.S. EPA, 2009). Because of this potentially increased access, it has been postulated that 

4 increases in epithelial permeability following O3 exposure might lead to increases in airway 

5 responsiveness to specific and nonspecific agents. In a recent study, Que et al. (2011) 

6 investigated this hypothesis in healthy young adults (83M, 55 F) exposed to 220 ppb O3 for 2.25 

7 hours (alternating 15 min periods of rest and brisk treadmill walking). As has been observed for 

8 FEV1 responses, within-individual changes in permeability were correlated between sequential 

9 O3 exposures, indicating intrinsic differences among individuals in susceptibility to epithelial 

10 damage following O3 exposures. However, increases in epithelial permeability at 1 day post-O3

11 exposure were not correlated with with changes in airway responsiveness assessed 1 day post-O3

12 exposure. The authors concluded that changes in epithelial permeability is relatively constant 

13 over time in young healthy adults, although changes in permeability and AHR appear to be 

14 mediated by different physiologic pathways. 

15 The limited epidemiologic evidence reviewed in the 2006 O3 AQCD (U.S. EPA, 2006) 

16 demonstrated an association between short-term increases in ambient O3 concentrations and 

17 airways inflammation in children (1-hour max O3 of approximately 100 ppb). In the 2006 O3

18 AQCD (U.S. EPA, 2006), there was limited evidence for increases in nasal lavage levels of 

19 inflammatory cell counts and molecules released by inflammatory cells (i.e., eosinophilic 

20 cationic protein, and myeloperoxidases). Since 2006, as a result of the development of less 

21 invasive methods, there has been a large increase in the number of studies assessing ambient O3-

22 associated changes in airway inflammation and oxidative stress, the types of biological samples 

23 collected (e.g., lower airway), and the types of indicators. Most of these recent studies have 

24 evaluated biomarkers of inflammation or oxidative stress in exhaled breath, nasal lavage fluid, or 

25 induced sputum (U.S. EPA, 2013, section 6.2.3.2). These recent studies form a larger database to 

26 establish coherence with findings from controlled human exposure and animal studies that have 

27 measured the same or related biological markers. Additionally, results from these studies provide 

28 further biological plausibility for the associations observed between ambient O3 concentrations 

29 and respiratory symptoms and asthma exacerbations. 

30 A number of epidemiologic studies provide evidence that short-term increases in ambient 

31 O3 exposure increase pulmonary inflammation and oxidative stress in children, including those 
3-21

1 with asthma (Sienra-Monge et al., 2004; Barraza-Villarreal et al., 2008; Romieu et al., 2008; 

Berhane et al., 2011). Multiple studies examined and found increases in exhaled NO (eNO)16 2

3 (Berhane et al., 2011; Khatri et al., 2009; Barraza-Villarreal et al., 2008). In some studies of 

4 subjects with asthma, increases in ambient O3 concentration at the same lag were associated with 

5 both increases in pulmonary inflammation and respiratory symptoms (Khatri et al., 2009; 

6 Barraza-Villarreal et al., 2008). Although more limited in number, epidemiologic studies also 

7 found associations with cytokines such as IL-6 or IL-8 (Barraza-Villarreal et al., 2008; Sienra8 Monge et al., 2004), eosinophils (Khatri et al., 2009), antioxidants (Sienra-Monge et al., 2004), 

9 and indicators of oxidative stress (Romieu et al., 2008) (ISA, Section 6.2.3.2, U.S. EPA, 2013). 

10 Because associations with inflammation were attenuated with higher antioxidant intake the study 

11 by Sienra-Monge et al. (2004) provides additional evidence that inhaled O3 is likely to be an 

12 important source of reactive oxygen species in airways and/or may increase pulmonary 

13 inflammation via oxidative stress-mediated mechanisms among all age groups. Limitations in 

14 some recent studies have contributed to inconsistent results in adults (U.S. EPA, 2013, section 

15 6.2.3.2).

16 Exposure to ambient O3 on multiple days can result in larger increases in pulmonary 

17 inflammation and oxidative stress, as discussed in section 6.2.3.2 of the ISA (U.S. EPA, 2013). 

18 In studies that examined multiple O3 lags, multiday averages of 8-hour maximum or 

19 8-hour average concentrations were associated with larger increases in pulmonary inflammation 

20 and oxidative stress (Berhane et al., 2011; Delfino et al., 2010a; Sienra-Monge et al., 2004), 

21 consistent with controlled human exposure (U.S. EPA, 2013, section 6.2.3.1) and animal studies 

22 (U.S. EPA, 2013, section 6.2.3.3) reporting that some markers of pulmonary inflammation 

23 remain elevated with O3 exposures repeated over multiple days. Evidence from animal 

24 toxicological studies also clearly indicates that O3 exposures result in damage and inflammation 

25 in the lung (ISA, Section 5.3, U.S. EPA, 2013). In the few studies that evaluated the potential for 

26 confounding, O3 effect estimates were not confounded by temperature or humidity, and were 

27 robust to adjustment for PM2.5 or PM10 (Barraza-Villarreal et al., 2008; Romieu et al., 2008; 

28 Sienra-Monge et al., 2004).

 

16 Exhaled NO has been shown to be a useful biomarker for airway inflammation in large population-based studies 

(Linn et al., 2009) (ISA, U.S. EPA, 2013, Section 7.2.4). 
3-22

1 In conclusion, a relatively small number of controlled human exposure studies evaluating 

2 O3-induced airway inflammation have become available since the last review. For purposes of 

3 reviewing the current O3 NAAQS, the most important of these recent studies reported a 

4 statistically significant increase in airway inflammation in healthy adults at moderate exertion 

5 following exposures to 60 ppb O3, the lowest concentration that has been evaluated for 

6 inflammation. In addition, a number of recent epidemiologic studies report O3-associated 

7 increases in markers of pulmonary inflammation, particularly in children. Thus, recent studies 

8 continue to support the evidence for airway inflammation and injury that was available in 

9 previous reviews, with new evidence for such effects following exposures to lower 

10 concentrations than had been evaluated previously. 

11 Airway Hyperresponsiveness 

12 Airway hyperresponsiveness (AHR) refers to a condition in which the conducting 

13 airways undergo enhanced bronchoconstriction in response to a variety of stimuli. Airway 

14 hyperresponsiveness is an important consequence of exposure to ambient O3 because its presence 

15 reflects a change in airway smooth muscle reactivity, and indicates that the airways are 

16 predisposed to narrowing upon inhalation of a variety of ambient stimuli including specific 

17 triggers (i.e., allergens) and nonspecific triggers (e.g., SO2, and cold air). People with asthma are 

18 generally more sensitive to bronchoconstricting agents than those without asthma, and the use of 

19 an airway challenge to inhaled bronchoconstricting agents is a diagnostic test in asthma. 

20 Standards for airway responsiveness testing have been developed for the clinical laboratory 

21 (ATS, 2000a), although variation in the methodology for administering the bronchoconstricting 

22 agent may affect the results (Cockcroft et al., 2005). There is a wide range of airway 

23 responsiveness in people without asthma, and responsiveness is influenced by a number of 

24 factors, including cigarette smoke, pollutant exposures, respiratory infections, occupational 

25 exposures, and respiratory irritants. Dietary antioxidants have been reported to attenuate O3-

26 induced bronchial hyperresponsiveness in people with asthma (Trenga et al., 2001). 

27 Evidence for airway hyperresponsiveness following O3 exposures is derived primarily 

28 from controlled human exposure and toxicological studies (U.S. EPA, 2013, section 6.2.2). 

29 Airway responsiveness is often quantified by measuring changes in pulmonary function 

30 following the inhalation of an aerosolized allergen or a nonspecific bronchoconstricting agent 
3-23

1 (e.g., methacholine), or following exposure to a bronchoconstricting stimulus such as cold air. In 

2 the last review, controlled human exposure studies of mostly adults (≥ 18 years of age) had 

3 shown that exposures to O3 concentrations at or above 80 ppb increase airway responsiveness, as 

4 indicated by a reduction in the concentration of specific (e.g., ragweed) and non-specific (e.g., 

5 methacholine) agents required to produce a given reduction in lung function (e.g., as measured 

6 by FEV1 or specific airway resistance) (U.S. EPA, 2013, section 6.2.2.1). This O3-induced AHR 

7 has been reported to be dose-dependent (Horstman et al., 1990). Animal toxicology studies have 

8 reported O3-induced airway hyperresponsiveness in a number of species, with some rat strains 

9 exhibiting hyperresponsiveness following 4-hour exposures to O3 concentrations as low as 50 

10 ppb (Depuydt et al., 1999). Since the last review, there have been relatively few new controlled 

11 human exposure and animal toxicology studies of O3 and airway hyperresponsiveness, and no 

12 new studies have evaluated exposures to O3 concentrations at or below 80 ppb (U.S. EPA, 2013, 

13 section 6.2.2.1) 

14 Airway hyperresponsiveness is linked with the accumulation and/or activation of 

15 eosinophils in the airways of asthmatics, which is followed by production of mucus and a late16 phase asthmatic response (75 FR 2970, January 19, 2010). In a study of 16 intermittent 

17 asthmatics, Hiltermann et al. (1999) found that there was a significant inverse correlation 

18 between the O3-induced change in the percentage of eosinophils in induced sputum and the 

19 concentration of methacholine causing a 20% decrease in FEV1. Hiltermann et al. (1999) 

20 concluded that the results point to the role of eosinophils in O3-induced airway 

21 hyperresponsiveness. Increases in O3-induced nonspecific airway responsiveness incidence and 

22 duration could have important clinical implications for children and adults with asthma, such as 

23 exacerbations of their disease. 

24 Airway hyperresponsiveness after O3 exposure appears to resolve more slowly than 

25 changes in FEV1 or respiratory symptoms (Folinsbee and Hazucha, 2000). Studies suggest that 

26 O3-induced AHR usually resolves 18 to 24 hours after exposure, but may persist in some 

27 individuals for longer periods (Folinsbee and Hazucha, 1989). Furthermore, in studies of 

28 repeated exposure to O3, changes in AHR tend to be somewhat less susceptible to attenuation 

29 with consecutive exposures than changes in FEV1 (Gong et al., 1997a; Folinsbee et al., 1994; 

30 Kulle et al., 1982; Dimeo et al., 1981) (U.S. EPA, 2013, section 6.2.2). In animal studies a 3-day 

31 continuous exposure resulted in attenuation of O3-induced airway hyperresponsiveness (Johnston 
3-24

1 et al., 2005) while repeated exposures for 2 hours per day over 10 days did not (Chhabra et al., 

2 2010), suggesting that attenuation could be lost when repeated exposures are interspersed with 

3 periods of rest (U.S. EPA, 2013, section 6.2.2.2). 

4 Increases in airway responsiveness do not appear to be strongly associated with 

5 decrements in lung function or increases in symptoms (Aris et al., 1995). Recently, Que et al. 

6 (2011) assessed methacholine responsiveness in healthy young adults (83M, 55 F) one day after 

7 exposure to 220 ppb O3 and filtered air for 2.25 hours (alternating 15 minute periods of rest and 

8 brisk treadmill walking). Increases in airways responsiveness at 1 day post-O3 exposure were not 

9 correlated with FEV1 responses immediately following the O3 exposure or with changes in 

10 epithelial permeability assessed 1-day post-O3 exposure. This indicates that airway hyper11 responsiveness also appears to be mediated by a differing physiologic pathway. 

12 As mentioned above, in addition to human subjects a number of species, including 

13 nonhuman primates, dogs, cats, rabbits, and rodents, have been used to examine the effect of O3

14 exposure on airway hyperresponsiveness (see Table 6-14, (U.S. EPA, 1996n) of the 1996 O3

15 AQCD and Annex Table AX5-12 on page AX5-36 (U.S. EPA, 2006h) of the 2006 O3 AQCD). A 

16 body of animal toxicology studies, including some recent studies conducted since the last review, 

17 provides support for the O3-induced AHR reported in humans (U.S. EPA, 2013, section 6.2.2.2). 

18 Although most of these studies evaluated O3 concentrations above those typically found in 

19 ambient air in cities in the United States (i.e., most studies evaluated O3 concentrations of 100 

20 ppb or greater), one study reported that a very low exposure concentration (50 ppb for 4 hours) 

21 induced AHR in some rat strains (Depuydt et al., 1999). Additional recent rodent studies 

22 reported O3-induced AHR following exposures to O3 concentrations from 100 to 500 ppb 

23 (Johnston et al., 2005; Chhabra et al., 2010; Larsen et al., 2010). In characterizing the relevance 

24 of these exposure concentrations, the ISA noted that a study using radiolabeled O3 suggests that 

25 even very high O3 exposure concentrations in rodents could be equivalent to much lower 

26 exposure concentrations in humans. Specifically, a 2000 ppb (2 ppm) O3 exposure concentration 

27 in resting rats was reported to be roughly equivalent to a 400 ppb exposure concentration in 

28 exercising humans (Hatch et al., 1994). Given this relationship, the ISA noted that animal data 

29 obtained in resting conditions could underestimate the risk of effects for humans (U.S. EPA, 

30 2013, section 2.4, p. 2-14). 
3-25

1 The 2006 AQCD (U.S. EPA, 2006, p. 6-34) concluded that spirometric responses to O3

2 are independent of inflammatory responses and markers of epithelial injury (Balmes et al., 1996; 

3 Blomberg et al., 1999; Hazucha et al., 1996; Torres et al., 1997). Significant inflammatory 

4 responses to O3 exposures that did not elicit significant spirometric responses have been reported 

5 (Holz et al., 2005; McBride et al., 1994). A recent study (Que et al., 2011) indicates that airway 

6 hyper-responsiveness also appears to be mediated by a differing physiologic pathway. These 

7 results from controlled human exposure studies indicate that sub-populations of healthy study 

8 subjects consistently experience larger than average lung function decrements, greater than 

9 average inflammatory responses and pulmonary injury as expressed by increased epithelial 

10 permeability, and greater than average airway responsiveness, and that these effects are mediated 

11 by apparently different physiologic pathways. Except for lung function decrements, we do not 

12 have the concentration- or exposure-response function information about the other, potentially 

more sensitive,17 13 clinical endpoints (i.e., inflammation, increased epithelial permeability, airway 

14 hyperresponsiveness) that would allow us to quantitatively estimate the size of the population 

15 affected and the magnitude of their responses. Moreover, some uncertainties about the exact 

16 physiological pathways underlying these endpoints prevents us from knowing whether the 

17 exaggerated responses are distributed in sub-populations evenly across the population, or may be 

18 clustered with more than one type of exaggerated response in particular sub-populations, or both. 

19 In summary, a strong body of controlled human exposure and animal toxicological 

20 studies, most of which were available in the last review of the O3 NAAQS, report O3-induced 

21 airway hyperresponsiveness after either acute or repeated exposures (U.S. EPA, 2013, section 

22 6.2.2.2). People with asthma often exhibit increased airway responsiveness at baseline relative to 

23 healthy controls, and they can experience further increases in responsiveness following 

24 exposures to O3. Studies reporting increased airway responsiveness after O3 exposure contribute 

25 to a plausible link between ambient O3 exposures and increased respiratory symptoms in 

26 asthmatics, and increased hospital admissions and emergency department visits for asthma (U.S. 

27 EPA, 2013, section 6.2.2.2). 

 

17 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

1 Respiratory Symptoms and Medication Use 

2 Because respiratory symptoms are associated with adverse outcomes such as limitations 

3 in activity, and are the primary reason for people with asthma to use quick relief medication and 

4 seek medical care, studies evaluating the link between O3 exposures and such symptoms allow a 

5 more direct characterization of the clinical and public health significance of ambient O3 exposure 

6 than measures of lung function decrements and pulmonary inflammation. Controlled human 

7 exposure and toxicological studies have described modes of action through which short-term O3

8 exposures may increase respiratory symptoms by demonstrating O3-induced airway 

9 hyperresponsiveness (U.S. EPA, 2013, section 6.2.2) and pulmonary inflammation (U.S. EPA, 

10 2013, section 6.2.3). 

11 The link between subjective respiratory symptoms and O3 exposures has been evaluated 

12 in both controlled human exposure and epidemiologic studies, and the link with medication use 

13 has been evaluated in epidemiologic studies. In the last review, several controlled human 

14 exposure studies reported respiratory symptoms following exposures to O3 concentrations at or 

15 above 80 ppb. In addition, one study reported such symptoms following exposures to 60 ppb O3, 

16 though the increase was not statistically different from filtered air controls. Epidemiologic 

17 studies reported associations between ambient O3 and respiratory symptoms and medication use 

18 in a variety of locations and populations, including asthmatic children living in U.S. cities. In the 

19 current review, additional controlled human exposure studies have evaluated respiratory 

20 symptoms following exposures to O3 concentrations below 80 ppb and recent epidemiologic 

21 studies have evaluated associations with respiratory symptoms and medication use (U.S. EPA, 

22 2013, sections 6.2.1, 6.2.4). 

23 In controlled human exposure studies available in the last review as well as newly 

24 available studies, statistically significant increases in respiratory symptoms have been 

25 consistently reported in healthy volunteers engaged in intermittent, moderate exertion following 

26 6.6 hour exposures to average O3 concentrations at or above 80 ppb (Adams, 2003; Adams, 

27 2006; Schelegle et al., 2009). Such symptoms have been reported to increase with increasing O3

28 exposure concentrations, duration of exposure, and activity level (McDonnell et al., 1999b). For 

29 example, in a study available during the last review, Adams (2006) reported an increase in 

30 respiratory symptoms in healthy adults during a 6.6 hour exposure protocol with an average O3
3-27

1 exposure concentration of 60 ppb. This increase was significantly different from initial 

2 respiratory symptoms, but not from filtered air controls. Two recent controlled human exposure 

3 studies that have become available since the last review did not report statistically significant 

4 increases in respiratory symptoms following exposures of healthy adults to 60 ppb O3 (Schelegle 

5 et al., 2009; Kim et al., 2011). A recent study by Schelegle et al. (2009) did report a statistically 

6 significant increase in respiratory symptoms in healthy adults following 6.6 hour exposures to an 

7 average O3 concentration of 70 ppb. The findings for O3-induced respiratory symptoms in 

8 controlled human exposure studies, and the evidence integrated across disciplines describing 

9 underlying modes of action, provide biological plausibility for epidemiologic associations 

10 observed between short-term increases in ambient O3 concentration and increases in respiratory 

11 symptoms (U.S. EPA, 2013, section 6.2.4). 

12 In epidemiologic studies, respiratory symptom data typically are collected by having 

13 subjects (or their parents) record symptoms and medication use in a diary without direct 

14 supervision by study staff. Several limitations of symptom reports are well recognized, as 

15 described in the ISA (U.S. EPA, 2013, section 6.2.4). Nonetheless, symptom diaries remain a 

16 convenient tool to collect individual-level data from a large number of subjects and allow 

17 modeling of associations between daily changes in O3 concentration and daily changes in 

18 respiratory morbidity over multiple weeks or months. Importantly, many of the limitations in 

19 these studies are sources of random measurement error that can bias effect estimates to the null 

20 or increase the uncertainty around effect estimates (U.S. EPA, 2013, Section 6.2.4). Because 

21 respiratory symptoms are associated with limitations in activity and daily function and are the 

22 primary reason for using medication and seeking medical care, the evidence is directly coherent 

23 with the associations consistently observed between increases in ambient O3 concentration and 

24 increases in asthma emergency department visits, discussed below (U.S. EPA, 2013, Section 

25 6.2.4). 

26 Most epidemiologic studies of O3 and respiratory symptoms and medication use have 

27 been conducted in children and/or adults with asthma, with fewer studies, and less consistent 

28 results, in non-asthmatic populations (U.S. EPA, 2013, section 6.2.4). The 2006 AQCD (U.S. 

29 EPA, 2006b, U.S. EPA, 2013, section 6.2.4) concluded that the collective body of epidemiologic 

30 evidence indicated that short-term increases in ambient O3 concentrations are associated with 

31 increases in respiratory symptoms in children with asthma. A large body of single-city and 
3-28

1 single-region studies of asthmatic children provides consistent evidence for associations between 

2 short-term increases in ambient O3 concentrations and increased respiratory symptoms and 

3 asthma medication use in children with asthma (U.S. EPA, 2013, Figure 6-12, Table 6-20). 

4 Methodological differences among studies make comparisons across recent multicity 

5 studies of respiratory symptoms difficult. Because of fewer person-days of data (Schildcrout et 

6 al., 2006) or examination of 19-day averages of ambient O3 concentrations (O'Connor et al., 

7 2008), the ISA did not give greater weight to results from recent multicity studies than results 

8 from single-city studies (U.S. EPA, 2013, section 6.2.4.5). While evidence from the few 

9 available U.S. multicity studies is less consistent (O'Connor et al., 2008; Schildcrout et al., 2006; 

10 Mortimer et al., 2002), the overall body of epidemiologic evidence with respect to the 

11 association betweeen exposure to O3 and respiratory symptoms in asthmatic children remains 

12 compelling (U.S. EPA, 2013, section 6.2.4.1). Findings from a small body of studies indicate that 

13 O3 is also associated with increased respiratory symptoms in adults with asthma (Khatri et al., 

14 2009; Feo Brito et al., 2007; Ross et al., 2002) (U.S. EPA, 2013, section 6.2.4.2). 

15 Available evidence indicates that O3-associated increases in respiratory symptoms are not 

16 confounded by temperature, pollen, or copollutants (primarily PM) (U.S. EPA, 2013, section 

17 6.2.4.5; Table 6-25; Romieu et al., 1996; Romieu et al., 1997; Thurston et al., 1997; Gent et al., 

18 2003). However, identifying the independent effects of O3 in some studies was complicated due 

19 to the high correlations observed between O3 and PM or different lags and averaging times 

20 examined for copollutants. Nonetheless, the ISA noted that the robustness of associations in 

21 some studies of individuals with asthma, combined with findings from controlled human 

22 exposure studies for the direct effects of O3 exposure, provide substantial evidence supporting 

23 the independent effects of short-term ambient O3 exposure on respiratory symptoms (U.S. EPA, 

24 2013, section 6.2.4.5). 

25 Epidemiologic studies of medication use have reported associations with 

26 1-hour maximum O3 concentrations and with multiday average O3 concentrations (Romieu et al.,

27 2006; Just et al., 2002). Some studies reported O3 associations for both respiratory symptoms and 

28 asthma medication use (Escamilla-Nuñez et al., 2008; Romieu et al., 2006; Schildcrout et al., 

29 2006; Jalaludin et al., 2004; Romieu et al., 1997; Thurston et al., 1997) while others reported 

30 associations for either respiratory symptoms or medication use (Romieu et al., 1996; Rabinovitch 

31 et al., 2004; Just et al., 2002; Ostro et al., 2001). 

32 In summary, both controlled human exposure and epidemiologic studies have reported 

33 respiratory symptoms attributable to short-term O3 exposures. In the last review, the majority of 

34 the evidence from controlled human exposure studies in young, healthy adults was for symptoms 
3-29

1 following exposures to O3 concentrations at or above 80 ppb. Although studies that have become 

2 available since the last review have not reported respiratory symptoms in young, healthy adults 

3 following exposures with moderate exertion to 60 ppb, one recent study has reported increased 

4 symptoms in young, healthy adults while at moderate exertion following exposures to O3

5 concentrations as low as 70 ppb. As was concluded in the 2006 O3 AQCD (U.S. EPA, 2006b, 

6 1996a), the collective body of epidemiologic evidence indicates that short-term increases in 

7 ambient O3 concentration are associated with increases in respiratory symptoms in children with 

8 asthma (U.S. EPA, 2013, section 6.2.4). Recent studies of respiratory symptoms and medication 

9 use, primarily in asthmatic children, add to this evidence. In a smaller body of studies, increases 

10 in ambient O3 concentration were associated with increases in respiratory symptoms in adults 

11 with asthma. 

12 Lung Host Defense 

13 The mammalian respiratory tract has a number of closely integrated defense mechanisms 

14 that, when functioning normally, provide protection from the potential health effects of 

15 exposures to a wide variety of inhaled particles and microbes. These defense mechanisms 

include mucociliary clearance, alveolobronchiolar transport mechanism, alveolar macrophages18 16 , 

and adaptive immunity19 17 (U.S. EPA, 2013, section 6.2.5). The previous O3 AQCD (U.S. EPA, 

18 2006) concluded that animal toxicological studies provided evidence that acute exposure to O3

19 concentrations as low as 100 to 500 ppb can increase susceptibility to infectious diseases due to 

20 modulation of these lung host defenses. This conclusion was based in large part on animal 

21 studies of alveolar macrophage functioning and mucociliary clearance (U.S. EPA, 2013, section 

22 6.2.5). 

23 With regard to alveolar macrophage functioning, the previous O3 AQCD (U.S. EPA, 

24 2006) concluded that short periods of O3 exposure can cause a reduction in the number of free 

25 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 lysosomal20 26

27 enzyme activities (U.S. EPA, 2013, section 6.2.5). These conclusions were based largely on 

 

18 Phagocytic white blood cells within the alveoli of the lungs that ingest inhaled particles. 

19 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. 

20 Lysosomes are cellular organelles that contain acid hydrolase enzymes that break down waste materials and 

cellular debris. 
3-30

1 studies conducted in animals exposed for several hours up to several weeks to O3 concentrations 

2 from 100 to 250 ppb (Hurst et al., 1970; Driscoll et al., 1987; Cohen et al., 2002). Consistent 

3 with the animal evidence, a controlled human exposure study available in the last review had 

4 reported decrements in the ability of alveolar macrophages to phagocytize yeast following 

5 exposures of healthy volunteers to O3 concentrations of 80 and 100 ppb for 6.6-hour during 

6 moderate exercise (Devlin et al., 1991). Integrating the animal study results with human 

7 exposure evidence available in the 1996 Criteria Document, the 2006 Criteria Document 

8 concluded that available evidence indicates that short-term O3 exposures have the potential to 

9 impair host defenses in humans, primarily by interfering with alveolar macrophage function. Any 

10 impairment in alveolar macrophage function may lead to decreased clearance of microorganisms 

11 or nonviable particles. Compromised alveolar macrophage functions in asthmatics may increase 

12 their susceptibility to other O3 effects, the effects of particles, and respiratory infections (EPA, 

13 2006a, p. 8–26). 

14 With regard to mucociliary clearance, in the last review a number of studies conducted in 

15 different animal species had reported morphological damage to the cells of the tracheobronchial 

16 tree from acute and sub-chronic exposure to O3 concentrations at or above 200 ppb. The cilia 

17 were either completely absent or had become noticeably shorter or blunt. In general, functional 

18 studies of mucociliary transport had observed a delay in particle clearance soon after acute 

19 exposure, with decreased clearance more evident at higher doses (1 ppm) (U.S. EPA, 1986; U.S. 

20 EPA, 2013, section 6.2.5.1).

21 Alveolobronchiolar transport mechanisms refers to the transport of particles deposited in 

22 the deep lung (alveoli) which may be removed either up through the respiratory tract (bronchi) 

23 by alveolobronchiolar transport or through the lymphatic system. The pivotal mechanism of 

24 alveolobronchiolar transport involves the movement of alveolar macrophages with ingested 

25 particles to the bottom of the conducting airways. These airways are lined with ciliated epithelial 

26 cells and cells that produce mucous, which surrounds the macrophages. The ciliated epithelial 

27 cells move the mucous packets up the resiratory tract, hence the term “mucociliary escalator.” 

28 Although some studies show reduced tracheobronchial clearance after O3 exposure, alveolar 

29 clearance of deposited material is accelerated, presumably due to macrophage influx, which in 

30 itself can be damaging. 
3-31

1 With regard to adaptive immunity, a limited number of epidemiologic studies have 

2 examined associations between O3 exposure and hospital admissions or ED visits for respiratory 

3 infection, pneumonia, or influenza. Results have been mixed, and in some cases conflicting (U.S. 

4 EPA, 2013, Sections 6.2.7.2 and 6.2.7.3). With the exception of influenza, it is difficult to 

5 ascertain whether cases of respiratory infection or pneumonia are of viral or bacterial etiology. 

6 A recent study that examined the association between O3 exposure and respiratory hospital 

7 admissions in response to an increase in influenza intensity did observe an increase in respiratory 

8 hospital admissions (Wong et al., 2009), but information from toxicological studies of O3 and 

9 viral infections is ambiguous. 

10 In summary, relatively few studies conducted since the last review have evaluated the 

11 effects of O3 exposures on lung host defense. When the available evidence is taken as a whole, 

12 the ISA concludes that acute O3 exposures impair the host defense capability of animals, 

13 primarily by depressing alveolar macrophage function and perhaps also by decreasing 

14 mucociliary clearance of inhaled particles and microorganisms. Coupled with limited evidence 

15 from controlled human exposure studies, this suggests that humans exposed to O3 could be 

16 predisposed to bacterial infections in the lower respiratory tract (EPA, 2013, section 6.2.5.5).

17 The seriousness of such infections may depend on how quickly bacteria develop virulence 

18 factors and how rapidly PMNs are mobilized to compensate for the deficit in alveolar 

19 macrophage function. 

20 Allergic and Asthma-Related Responses 

21 Effects resulting from combined exposures to O3 and allergens have been studied in a 

22 variety of animal species, generally as models of experimental asthma. Pulmonary function and 

23 AHR in animal models of asthma are discussed in detail in Section 6.2.1.3 and Section 6.2.2.2, 

24 respectively, in the ISA (U.S. EPA, 2013). Studies of allergic and asthma-related responses are 

25 discussed in detail in sections 5.3.6 and 6.2.6 of the ISA (U.S. EPA, 2013). 

26 Evidence available in the last review indicates that O3 exposure skews immune responses 

27 toward an allergic phenotype. For example, Gershwin et al. (1981) reported that O3 (800 and 500 

28 ppb for 4 days) exposure caused a 34-fold increase in the number of IgE (allergic antibody)-

29 containing cells in the lungs of mice. In general, the number of IgE-containing cells correlated 

30 positively with levels of anaphylactic sensitivity. In humans, allergic rhinoconjunctivitis 
3-32

1 symptoms are associated with increases in ambient O3 concentrations (Riediker et al., 2001). 

2 Controlled human exposure studies have observed O3-induced changes indicating allergic 

3 skewing. Airway eosinophils, which are white blood cells that participate in allergic disease and 

inflammation, were observed to increase in volunteers with atopy21 4 and mild asthma (Peden et 

5 al., 1997). In a more recent study, expression of IL-5, a cytokine involved in eosinophil 

6 recruitment and activation, was increased in subjects with atopy but not in healthy subjects 

7 (Hernandez et al., 2010). Epidemiologic studies describe associations between eosinophils in 

8 both short- (U.S. EPA, 2013, Section 6.2.3.2) and long-term (U.S. EPA, 2013, Section 7.2.5) O3

9 exposure, as do chronic exposure studies in non-human primates. Collectively, findings from 

10 these studies suggest that O3 can induce or enhance certain components of allergic inflammation 

11 in individuals with allergy or allergic asthma. 

12 Evidence available in the last review indicates that ozone may also increase AHR to 

13 specific allergen triggers (75 FR 2970, January 19, 2010). Two studies (Jörres et al., 1996; Holz 

14 et al., 2002) observed increased airway responsiveness to O3 exposure with bronchial allergen 

15 challenge in subjects with preexisting allergic airway disease. Ozone-induced exacerbation of 

16 airway responsiveness persists longer and attenuates more slowly than O3-induced lung function 

17 decrements and respiratory symptom responses and can have important clinical implications for 

18 asthmatics. 

19 Animal toxicology studies indicate that O3 enhances inflammatory and allergic responses 

20 to allergen challenge in sensitized animals. In addition to exacerbating existing allergic 

21 responses, toxicology studies indicate that O3 can also act as an adjuvant to produce sensitization 

22 in the respiratory tract. Along with its pro-allergic effects (inducing or enhancing certain 

23 components of allergic inflammation in individuals with allergy or allergic asthma), O3 could 

24 also make airborne allergens more allergenic. When combined with NO2, O3 has been shown to 

25 enhance nitration of common protein allergens, which may increase their allergenicity Franze et 

26 al. (2005). 

 

21 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. 
3-33

1 Hospital Admissions and Emergency Department Visits 

2 The 2006 O3 AQCD evaluated numerous studies of respiratory-related emergency 

3 department visits and hospital admissions. These were primarily time-series studies conducted in 

4 the U.S., Canada, Europe, South America, Australia, and Asia. Based on such studies, the 2006 

5 O3 AQCD concluded that “the overall evidence supports a causal relationship between acute 

6 ambient O3 exposures and increased respiratory morbidity resulting in increased ED visits and 

[hospital admissions] during the warm season22 7 ” (U.S. EPA, 2006). This conclusion was 

8 “strongly supported by the human clinical, animal toxicologic[al], and epidemiologic evidence 

9 for [O3-induced] lung function decrements, increased respiratory symptoms, airway 

10 inflammation, and airway hyperreactivity” (U.S. EPA, 2006). 

11 The results of recent studies largely support the conclusions of the 2006 O3AQCD (U.S. 

12 EPA, 2013, section 6.2.7). Since the completion of the 2006 O3 AQCD, relatively fewer studies 

13 conducted in the U.S., Canada, and Europe have evaluated associations between short-term O3

14 concentrations and respiratory hospital admissions and emergency department visits, with a 

15 growing number of studies conducted in Asia. This epidemiologic evidence is summarized in 

16 Appendix 3-B and discussed in detail in the ISA (U.S. EPA, 2013, section 6.2.7). 

17 In considering this body of evidence, the ISA focused primarily on multicity studies 

18 because they examine associations with respiratory-related hospital admissions and emergency 

19 department visits over large geographic areas using consistent statistical methodologies (U.S. 

20 EPA, 2013, section 6.2.7.1). The ISA also focused on single-city studies that encompassed a 

21 large number of hospital admissions or emergency department visits, included long study22 durations, were conducted in locations not represented by the larger studies, or examined 

23 population-specific characteristics that were not evaluated in the larger studies (U.S. EPA, 2013, 

24 section 6.2.7.1). When examining the association between short-term O3 exposure and 

25 respiratory health effects that require medical attention, the ISA distinguishes between hospital 

26 admissions and emergency department visits because it is likely that a small percentage of 

27 respiratory emergency department visits will be admitted to the hospital, therefore respiratory 

28 emergency department visits may represent potentially less serious, but more common outcomes 

29 (U.S. EPA, 2013, section 6.2.7.1). 

30 The results of studies evaluated in this review largely support the conclusions of the 2006 

31 AQCD. Several recent multicity studies (e.g., Cakmak et al., 2006; Dales et al., 2006) and a 

32 multi-continent study (Katsouyanni et al., 2009) report associations between short-term O3

 

22Epidemiologic 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. 
3-34

1 concentrations and increased respiratory-related hospital admissions and emergency department 

2 visits. These multicity studies are supported by consistent results from single-city studies using 

3 different exposure assignment approaches (i.e., average of multiple monitors, single monitor, 

4 population-weighted average) and averaging times (i.e., 1-hour max and 8-hour max) (U.S. EPA, 

5 2013, sections 6.2.7.1 to 6.2.7.5). Recent studies also report associations with hospital 

6 admissions and emergency department visits for asthma (Strickland et al., 2010; Stieb et al., 

7 2009) and COPD (Stieb et al., 2009; Medina-Ramon et al., 2006), with more limited evidence for 

8 pneumonia (Medina-Ramon et al., 2006; Zanobetti and Schwartz, 2006). In seasonal analyses 

9 (Figure 3-2 below; U.S. EPA, 2013, Figure 6-19, Table 6-28), stronger associations were 

10 reported in the warm season or summer months (red circles) compared to the cold season (blue 

11 circles), particularly for asthma (Strickland et al., 2010; Ito et al., 2007b) and COPD (Medina12 Ramon et al., 2006). The available evidence indicates that children are at greatest risk for 

13 O3-induced respiratory effects (Silverman and Ito, 2010; Strickland et al., 2010; Mar and Koenig, 

14 2009; Villeneuve et al., 2007; Dales et al., 2006). 

15 Although the collective evidence across studies indicates a mostly consistent positive 

16 association between O3 exposure and respiratory-related hospital admissions and ED visits, the 

17 magnitude of these associations may be underestimated due to behavioral modification in 

18 response to air quality forecasts (U.S. EPA, 2013, Section 4.6.6).
3-35

Note: Effect estimates are for a 20 ppb increase in 24-h; 30 ppb increase in 8-h max; and 40 ppb increase in 1-h max O3 concentrations. HA=hospital admission; 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. 

1

2 Figure 3-2. Percent increase in respiratory-related hospital admission and emergency department visits in studies that 

3 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)

Darrow et al. (2009)

Tolbert et 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-Canada

APHENA-Canada

Atlanta

Atlanta

8 Italian cities

APHENA-Europe

APHENA-U.S.

APHENA-Canada

APHENA-Canada

7 Canadian Cities

Alberta, CAN

Atlanta

New York

New 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

All

All

0-27 days 

0-2

65+

65+

65+

65+

All 

All

All

65+

65+

65+

65+

All

> 2

Children

All

All

> 2

18+

Children

6-1 8

<18

All

> 2

Children

All

All

65+

65+

All

65+

65+

65+

65+

65+

65+

Respiratory

Asthma

COPD

Pneumonia

◄

►

►

►

►

►

Lag

0-1

1.2

2

0-6

0-1

0-1

DL (0-2 )

DL (0-2)b

1

0-2

0-3

0-1

0-1

DL (0-2 )

DL (0-2)b

2

0-2

0-2

0-1

0-1

0-2

2

0-2

0-1

0

0-1

0-2

0-2

0-1

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

HA

HA

HA

ED

HA

HA

HA

HA

ED

ED

HA

HA

HA

HA

HA

ED

ED

ED

HA

ED

ED

ED

ED

HA

ED

ED

ED

ED

HA

ED

HA

HA

ED

HA

HA

HA

HA

HA

HA

-2 5 -2 0 -1 5 -1 0 -5 0 5 1 0 1 5 2 0 2 5 3 0 3 5 4 0

% Increase
3-36

1 Studies examining the potential confounding effects of copollutants have reported that O3

2 effect estimates remained relatively robust upon the inclusion of PM and gaseous pollutants in 

3 two-pollutant models (U.S. 2013, Figure 6-20, Table 6-29). Additional studies that conducted 

4 copollutant analyses, but did not present quantitative results, also support these conclusions 

5 (Strickland et al., 2010; Tolbert et al., 2007; Medina-Ramon et al., 2006) (U.S. 2013, section 

6 6.2.7.5). 

7 In the last review, studies had not evaluated the concentration-response relationship 

8 between short-term O3 exposure and respiratory-related hospital admissions and emergency 

9 department visits. A preliminary examination of this relationship in studies that have become 

10 available since the last review found no evidence of a deviation from linearity when examining 

11 the association between short-term O3 exposure and asthma hospital admissions (U.S. EPA, 

12 2013, page 6-157, and Silverman and Ito, 2010). In addition, an examination of the 

13 concentration-response relationship for O3 exposure and pediatric asthma emergency department 

14 visits found no evidence of a threshold at O3 concentrations as low as 30 ppb (for 8-hour daily 

15 maximum concentrations) (Strickland et al., 2010). However, in both studies there is uncertainty 

16 in the shape of the concentration-response curve at the lower end of the distribution of O3

17 concentrations due to the low density of data in this range (U.S. 2013, page 6-157).

18 Respiratory Mortality 

19 The controlled human exposure, epidemiologic, and toxicological studies discussed in 

20 section 6.2 of the ISA (U.S. EPA, 1013a, section 6.2) provide strong evidence for respiratory 

21 morbidity effects, including ED visits and hospital admissions, in response to short-term O3

22 exposures. Moreover, evidence from experimental studies indicates multiple potential pathways 

23 of respiratory effects from short-term O3 exposures, which support the continuum of respiratory 

24 effects that could potentially result in respiratory-related mortality in adults (U.S. EPA, 1013a, 

25 section 6.2.8). The 2006 O3 AQCD found inconsistent evidence for associations between short26 term O3 concentrations and respiratory mortality (U.S. EPA, 2006). Although some studies 

27 reported a strong positive association between O3 and respiratory mortality, additional studies 

28 reported small associations or no associations. New epidemiologic evidence for respiratory 

29 mortality is discussed in detail in section 6.2.8 of the ISA (U.S. EPA, 2013). The majority of 

30 recent multicity studies have reported positive associations between short-term O3 exposures and 

31 respiratory mortality, particularly during the summer months (U.S. EPA, 2013, Figure 6-36). 

32 Specifically, recent multicity studies from the U.S. (Zanobetti and Schwartz, 2008b), 

33 Europe (Samoli et al., 2009), Italy (Stafoggia et al., 2010), and Asia (Wong et al., 2010), as well 

34 as a multi-continent study (Katsouyanni et al., 2009), reported associations between short-term 
3-37

1 O3 concentrations and respiratory mortality (U.S. EPA, 2013, Figure 6-37, page 6-259). With 

2 respect to respiratory mortality, summer-only analyses were consistently positive and most were 

3 statistically significant. All-year analyses had more mixed results, but most were positive. 

4 Of the studies evaluated, only the studies by Katsouyanni et al. (2009) and by Stafoggia 

5 et al. (2010) analyzed the potential for copollutant confounding of the O3-respiratory mortality 

6 relationship. Based on the results of these analyses, the ISA concluded that O3 respiratory 

7 mortality risk estimates appear to be moderately to substantially sensitive (e.g., increased or 

8 attenuated) to inclusion of PM10. However, in the APHENA study, the mostly every-6th-day 

9 sampling schedule for PM10 in the Canadian and U.S. datasets greatly reduced their sample size 

10 and limits the interpretation of these results (U.S. EPA, 2013, section 6.2.8). 

11 In summary, recent epidemiologic studies support and reinforce the epidemiologic 

12 evidence for O3-associated respiratory hospital admissions and emergency department visits 

13 from the last review. In addition, the evidence for associations with respiratory mortality has 

14 been strengthened considerably since the last review, with the addition of several large multicity 

15 studies. The plausibility of the associations reported in these studies is supported by the 

16 experimental evidence for respiratory effects. 

17 3.1.2.2 Respiratory Effects – Long-term Exposures 

18 To what extent does the currently available scientific evidence, including related 

19 uncertainties, strengthen or alter our understanding from the last review of 

20 respiratory effects attributable to long-term O3 exposures?

21 As recognized in section 3.1.2.1, “the clearest evidence for health effects associated with 

22 exposure to O3 is provided by studies of respiratory effects” (U.S. EPA, 2013, section 1, p. 1-6). 

23 Collectively, there is a vast amount of evidence spanning several decades that supports a causal 

24 association between exposure to O3 and a continuum of respiratory effects (U.S. EPA, 2013, 

25 section 2.5). While the majority of this evidence is derived from studies investigating short-term 

26 exposures, evidence from animal toxicological studies and recent epidemiologic evidence 

27 indicate that long-term exposures (i.e., months to years) may also be detrimental to the 

28 respiratory system. Across this evidence, particularly the epidemiologic evidence, the exposures 

29 of focus vary, often involving repeated short concentrations extending over a long period, rather 

30 than a continuous long-term exposure period. 

31 In the 2006 O3 AQCD, evidence was examined for relationships between long-term O3

32 exposure and effects on respiratory health outcomes including declines in lung function, 

33 increases in inflammation, and development of asthma in children and adults. Animal toxicology 

34 data provided a clearer picture indicating that long-term O3 exposure may have lasting effects. 
3-38

Chronic23 1 exposure studies in animals have reported biochemical and morphological changes 

2 suggestive of irreversible long-term O3 impacts on the lung. In contrast to supportive evidence 

3 from chronic animal studies, the epidemiologic studies on longer-term (annual) lung function 

4 declines, inflammation, and new asthma development remained inconclusive. 

5 Several epidemiologic studies collectively indicated that O3 exposure averaged over 

6 several summer months was associated with smaller increases in lung function growth in 

7 children. For longer averaging periods (annual), the analysis in the Children’s Health Study 

8 (CHS) reported by Gauderman et al. (2004) provided little evidence that such long-term 

9 exposure to ambient O3 was associated with significant deficits in the growth rate of lung 

10 function in children. Limited epidemiologic research examined the relationship between long11 term O3 exposures and inflammation. Cross-sectional studies detected no associations between 

12 long-term O3 exposures and asthma prevalence, asthma-related symptoms or allergy to common 

13 aeroallergens in children. However, longitudinal studies provided evidence that long-term O3

14 exposure influences the risk of asthma development in children and adults. 

15 The currently available body of evidence supporting a relationship between long-term O3

16 exposures and adverse respiratory health effects that is likely to be causal is discussed in detail in 

17 the ISA (EPA 2013, section 7.2). New evidence reports interactions between genetic variants and 

18 long-term O3 exposure affect the occurrence of new-onset asthma in multi-community, U.S. 

19 cohort studies where protection by specific oxidant gene variants was restricted to children living 

20 in low O3 communities. A new line of evidence reports a positive concentration-response 

21 relationship between first asthma hospitalization and long-term O3 exposure. Related studies 

22 report coherent relationships between asthma severity and control, and respiratory symptoms 

23 among asthmatics and long-term O3 exposure. There is also limited evidence for an association 

24 between long-term exposure to ambient O3 concentrations and respiratory mortality. These 

25 studies are summarized briefly below for new-onset asthma and asthma prevalence, asthma 

26 hospital admissions and other morbidity effects, pulmonary structure and function, and 

27 respiratory mortality. 

28 Currently available scientific evidence of the adverse health effects attributable to long29 term O3 exposures, even considering related uncertainties, is much stronger than the body of 

30 evidence available at the time of the 2008 review of the O3 standard. The 2006 O3 AQCD (U.S. 

31 EPA 2006) concluded that epidemiologic studies provided no evidence of associations between 

32 long-term (annual) O3 exposures and asthma-related symptoms, asthma prevalence, or allergy to 

 

23 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. 
3-39

1 common allergens after controlling for covariates. It found limited evidence for a relationship 

2 between long-term exposures to ambient O3 and deficits in the growth rate of lung-function in 

3 children, pulmonary inflammation and other endpoints. Episodic exposures were also known to 

4 cause more severe pulmonary morphological changes than continuous exposure. 

5 The evidence base available in this review includes additional epidemiologic studies 

6 using a variety of designs and analysis methods evaluating the relationship between long-term O3

7 exposures and measures of respiratory morbidity and mortality effects conducted by different 

8 research groups in different locations. The ISA (U.S. EPA, 2013, p. 7-33), in Table 7-2 presents 

9 selected key new longitudinal and cross-sectional studies of respiratory health effects and 

10 associated O3 concentrations. The positive results from various designs and locations support a 

11 relationship between long-term exposure to ambient O3 and respiratory health effects and 

12 mortality. 

13 In this review, the evidence of effects associated with long-term exposures strengthens 

14 the relationship between O3 exposure and health effects defined as adverse by the American 

15 Thoracic Society (ATS), a definition that has been used in previous reviews of the O3 standard. 

16 As discussed in more detail in section 3.1.3 below, the ATS (1985) defined adverse as 

17 “medically significant physiologic or pathologic changes generally evidenced by one or more of 

18 the following: (1) interference with the normal activity of the affected person or persons, (2) 

19 episodic respiratory illness, (3) incapacitating illness, (4) permanent respiratory injury, and/or (5) 

20 progressive respiratory dysfunction.” As discussed below, in this review there is now credible 

21 evidence of respiratory health effects associated with long-term O3 exposures that would fall in 

22 to each of these five categories that define adversity. 

23 From a policy perspective, the recent epidemiologic studies from the CHS of long-term 

24 O3 exposures that shed light on the interaction between genetic variability, O3 exposures, and 

25 health effects in children are important, not only because they help clarify previous findings, but 

26 also because the effects evaluated, such as new-onset asthma, are clearly adverse. The ISA (U.S. 

27 EPA, 2013, p. 7-12) notes that the collective evidence from CHS provides an important 

28 demonstration of gene-environment interactions. It further notes that in the complex 

29 gene-environment setting a modifying effect might not be reflected in an exposure main effect 

30 and that the simultaneous occurrence of main effect and interaction effect can occur. Moreover, 

31 the study of gene-environment interactions elucidates disease mechanisms in humans by using 

32 information on susceptibility genes to focus on the biological pathways that are most relevant to 

33 that disease (Hunter, 2005).

34 In the CHS cohort of children in 12 Southern California communities, long-term 

35 exposure to O3 concentrations was not associated with increased risk of developing asthma 
3-40

1 (McConnell et al., 2010); however, greater outdoor exercise was associated with development of 

2 asthma in children living in communities with higher ambient O3 concentrations (McConnell et 

3 al., 2002). Recent CHS studies examined interactions among genetic variants, long-term O3

4 exposure, and new onset asthma in children. These prospective cohort studies are 

5 methodologically rigorous epidemiologic studies, and evidence indicates gene-O3 interactions. 

6 These studies have provided data supporting decreased risk of certain genetic variants on new 

7 onset asthma (e.g., HMOX-1, ARG) that is limited to children either in low (Islam et al., 2008) 

8 or high (Salam et al., 2009) O3 communities. Gene-environment interaction also was 

9 demonstrated with findings that greater outdoor exercise increased risk of asthma in GSTP1 

10 Ile/Ile children living in high O3 communities (Islam et al., 2009). Biological plausibility for 

11 these gene-O3 environment interactions is provided by evidence that these enzymes have 

12 antioxidant and/or anti-inflammatory activity and participate in well recognized modes of action 

13 in asthma pathogenesis. As O3 is a source of oxidants in the airways, oxidative stress serves as 

14 the link among O3 exposure, enzyme activity, and asthma. Cross-sectional studies by Akinbami 

15 et al. (2010) and Hwang et al. (2005) provide further evidence relating O3 exposures with asthma 

16 prevalence. 

17 Studies using a cross-sectional design provide support for a relationship between long18 term O3 exposure and adverse health effects in asthmatics, including: bronchitic symptoms 

19 (related to TNF-308 genotype in asthmatic children) (Lee et al., 2009b); asthma severity (Rage et 

20 al., 2009b) and asthma control (Jacquemin et al., 2012) in an adult cohort; respiratory-related 

21 school absences (related to CAT and MPO variant genes) (Wenten et al., 2009); asthma ED 

22 visits in adults (Meng et al., 2010); and, asthma hospital admissions in adults and children (Lin et 

23 al., 2008; Meng et al., 2010; Moore et al., 2008). Several studies, shown in Table 7-3 (ISA, U.S. 

24 EPA, 2013, p. 7-35), provide results adjusted for potential confounders presenting results for 

25 both O3 and PM (in single and multipollutant models) as well as other pollutants where PM 

26 effects were not provided. As shown in this table, O3 associations were generally robust to 

27 adjustment by potential confounding by PM. 

28 Information from toxicological studies in nonhuman primates indicates that long term 

29 exposure to O3 during gestation or development can result in irreversible morphological changes 

30 in the lung, which in turn can influence the function of the respiratory tract. This nonhuman 

31 primate evidence of an O3-induced change in airway responsiveness supports the biologic 

32 plausibility of long term exposure to O3 contributing to effects of asthma in children. However, 

33 results from epidemiologic studies examining long-term O3 exposure and pulmonary function 

34 effects are inconclusive with some new studies relating effects at higher exposure levels. 
3-41

1 The ISA (U.S. EPA, 2013, p. 7-31) concludes that there is limited evidence for an 

2 association between long-term exposure to ambient O3 concentrations and respiratory mortality 

3 in adults (Jerrett et al., 2009). This effect was robust to the inclusion of PM2.5 and insensitive to a 

4 number of different model specifications. Moreover, there is evidence that long-term exposure to 

5 O3 is associated with mortality among individuals that had previously experienced an emergency 

6 hospital admission due to COPD (Zanobetti and Schwartz, 2011). 

7 In conclusion, since the last review, the body of evidence about the effects of long-term 

8 O3 exposure has been considerably strengthened. The scientific evidence available for this 

9 review, including related uncertainties, provides an overall strong body of evidence of adverse 

10 health effects attributable to long-term O3 exposures. These include a coherent range of asthma 

11 morbidity effects such as new-onset asthma, asthma prevalence, symptoms, school absences, ED 

12 visits and hospital admissions. There is also new evidence of respiratory mortality associated 

13 with long-term O3 exposure. Further discussion of key studies is below. 

14 New-onset Asthma and Asthma Prevalence 

15 Asthma is a heterogeneous disease with a high degree of temporal variability. The on-set, 

16 progression, and symptoms can vary within an individual’s lifetime, and the course of asthma 

17 may vary markedly in young children, older children, adolescents, and adults. In the previous 

18 review, longitudinal cohort studies that examined associations between long-term O3 exposures 

19 and the onset of asthma in adults and children indicated a direct effect of long-term O3 exposures 

20 on asthma risk in adults (McDonnell et al., 1999a, 15-year follow-up; Greer et al., 1993, 10-year 

21 follow-up) and effect modification by O3 in children (McConnell et al., 2002). Since that review, 

22 important new evidence has become available about the association between long-term 

23 exposures to O3 and new-onset asthma that has increased our understanding of the gene24 environment interaction and the mechanisms and biological pathways most relevant to assessing 

25 O3-related effects. 

26 In children, the relationship between long-term O3 exposure and new-onset asthma has 

27 been extensively studied in the CHS; a long-term study that was initiated in the early1990’s 

28 which has evaluated effects in several cohorts of children. The CHS was initially designed to 

29 examine whether long-term exposure to ambient pollution was related to chronic respiratory 

30 outcomes in children in 12 communities in southern California. In the CHS, new-onset asthma 

31 was classified as having no prior history of asthma at study entry with subsequent report of 

32 physician-diagnosed asthma at follow-up, with the date of onset assigned to be the midpoint of 

33 the interval between the interview date when asthma diagnosis was first reported and the 

34 previous interview date. The results of one study (McConnell et al., 2002) available in the 
3-42

1 previous review indicated that within high O3 communities, asthma risk was 3.3 times greater for 

2 children who played three or more outdoor sports as compared with children who played no 

3 sports. 

4 For this review, as discussed in section 7.2.1.1 of the ISA (U.S. EPA 2013), recent 

5 studies from the CHS provide evidence for gene-environment interactions in effects on new6 onset asthma by indicating that the lower risks associated with specific genetic variants are found 

7 in children who live in lower O3 communities. These studies indicate that the risk for new-onset 

8 asthma is related in part to genetic susceptibility, as well as behavioral factors and environmental 

9 exposure. The onset of a chronic disease, such as asthma, is partially the result of a sequence of 

10 biochemical reactions involving exposures to various environmental agents metabolized by 

11 enzymes related to a number of different genes. Oxidative stress has been proposed to underlie 

12 the mechanistic hypotheses related to O3 exposure. Genetic variants may impact disease risk 

13 directly, or modify disease risk by affecting internal dose of pollutants and other environmental 

14 agents and/or their reaction products, or by altering cellular and molecular modes of action. 

15 Understanding the relation between genetic polymorphisms and environmental exposure can 

16 help identify high-risk subgroups in the population and provide better insight into pathway 

17 mechanisms for these complex diseases. 

18 The CHS analyses (Islam et al., 2008; Islam et al. 2009; Salam et al., 2009) have found 

19 that asthma risk is related to interactions between O3 and variants in genes for enzymes such as 

20 heme-oxygenase (HO-1), arginases (ARG1 and 2), and glutathione S transferase P1 (GSTP1). 

21 Biological plausibility for these findings is provided by evidence that these enzymes have 

22 antioxidant and/or anti-inflammatory activity and participate in well-recognized modes of action 

23 in asthma pathogenesis. Further, several lines of evidence demonstrate that secondary oxidation 

24 products of O3 initiate the key modes of action that mediate downstream health effects (ISA, 

25 Section 5.3, U.S. EPA, 2013). For example, HO-1 responds rapidly to oxidants, has anti26 inflammatory and anti-oxidant effects, relaxes airway smooth muscle, and is induced in the 

27 airways during asthma. Gene-environment interactions are discussed in detail in Section 5.4.2.1 

28 in the ISA (U.S. EPA, 2013). 

29 Asthma Hospital Admissions 

30 In the 2006 AQCD, studies on O3-related hospital discharges and emergency department 

31 (ED) visits for asthma and respiratory disease mainly looked at short-term (daily) metrics. The 

32 short-term O3 studies presented in section 6.2.7.5 of the ISA (U.S. EPA 2013) and discussed 

33 above in section 3.1.2.1 continue to indicate that there is evidence for increases in both hospital 

34 admissions and ED visits in children and adults related to all respiratory outcomes, including 
3-43

1 asthma, with stronger associations in the warm months. New studies, discussed in section 7.2.2 

2 of the ISA (U.S. EPA, 2013) also evaluated long-term O3 exposure metrics, providing a new line 

3 of evidence that suggests a positive exposure-response relationship between the first hospital 

4 admission for asthma and long-term O3 exposure, although the ISA cautions in attributing the 

5 associations in that study to long-term exposures since there is potential for short-term exposures 

6 to contribute to the observed associations. 

7 Evidence associating long-term O3 exposure to first asthma hospital admission in a 

8 positive concentration-response relationship is provided in a retrospective cohort study (Lin et 

9 al., 2008b). This study investigated the association between chronic exposure to O3 and 

10 childhood asthma admissions by following a birth cohort of more than 1.2 million babies born in 

11 New York State (1995-1999) to first asthma admission or until 31 December 2000. Three annual 

12 indicators (all 8-hour maximum from 10:00 a.m. to 6:00 p.m.) were used to define chronic O3

13 exposure: (1) mean concentration during the follow-up period (41.06 ppb); (2) mean 

14 concentration during the O3 season (50.62 ppb); and (3) proportion of follow-up days with O3

15 levels >70 ppb. The effects of co-pollutants were controlled, and interaction terms were used to 

16 assess potential effect modifications. A positive association between chronic exposure to O3 and 

17 childhood asthma hospital admissions was observed, indicating that children exposed to high O3

18 levels over time are more likely to develop asthma severe enough to be admitted to the hospital. 

19 The various factors were examined and differences were found for younger children (1-2 years), 

20 poor neighborhoods, Medicaid/self-paid births, geographic region and others. As shown in the 

21 ISA, Figure 7-3 (EPA 2013, p. 7-16), positive concentration-response relationships were 

22 observed. Asthma admissions were significantly associated with increased O3 levels for all 

23 chronic exposure indicators. 

24 In considering the relationship between long-term pollutant exposures and chronic 

25 disease heath endpoints, where chronic pathologies are found with acute expression of chronic 

26 disease, Künzli (2012) hypothesizes that if the associations of pollution with events are much 

27 larger in the long-term studies, it provides some indirect evidence that air pollution increases the 

28 pool of subjects with chronic disease, and that more acute events are to be expected to be seen 

29 for higher exposures. The results of Lin et al (2008) for first asthma hospital admission, 

30 presented in Figure 7-3 (EPA 2013, p. 7-16), show effects estimates that are larger than those 

31 reported in a study of childhood asthma hospital admission in New York state (Silverman and 

32 Ito, 2010), discussed in section 3.1.2.1 and 3.1.2.2 above. The ISA (U.S. EPA, 2013, p. 7-16) 

33 notes that this provides some support for the hypothesis that O3 exposure may not only have 

34 triggered the events but also increased the pool of asthmatic children, but cautions in attributing 
3-44

1 the associations in Lin et al. (2008) study to long-term exposures since there is potential for 

2 short-term exposures to contribute to the observed associations. 

3 Pulmonary structure and function 

4 In the 2006 O3 AQCD, few epidemiologic studies had investigated the effect of chronic 

5 O3 exposure on pulmonary function. The strongest evidence was for medium-term effects of 

6 extended O3 exposures over several summer months on lung function (FEV1) in children, i.e., 

7 reduced lung function growth being associated with higher ambient O3 levels. Short-term O3

8 exposure studies presented in ISA (EPA, 2013, Section 6.2.1.2), and above in section 3.1.2.1, 

9 provide a cumulative body of epidemiologic evidence that strongly supports associations 

10 between ambient O3 exposure and decrements in lung function among children. A recent study 

11 (Rojas-Martinez et al., 2007) of long-term exposure to O3, described in section 7.2.3.1 of the ISA 

12 (U.S. EPA, 2013, p. 7-19), observed a relationship with pulmonary function declines in school13 aged children where O3 and other pollutant levels were higher (90 ppb at high end of the range) 

14 than those in the CHS. Two studies of adult cohorts provide mixed results where long-term 

15 exposures were at the high end of the range. 

16 Long-term studies in animals allow for greater insight into the potential effects of 

17 prolonged exposure to O3 that may not be easily measured in humans, such as structural changes 

18 in the respiratory tract. Despite uncertainties, epidemiologic studies observing associations of O3

19 exposure with functional changes in humans can attain biological plausibility in conjunction with 

20 long-term toxicological studies, particularly O3-inhalation studies performed in non-human 

21 primates whose respiratory systems most closely resembles that of the human. An important 

22 series of studies, discussed in section 7.2.3.2 of the ISA (U.S. EPA, 2013), have used nonhuman 

23 primates to examine the effect of O3 alone, or in combination with an inhaled allergen, house 

24 dust mite antigen (HDMA), on morphology and lung function. These animals exhibit the 

25 hallmarks of allergic asthma defined for humans, including: a positive skin test for HDMA with 

26 elevated levels of IgE in serum and IgE-positive cells within the tracheobronchial airway walls; 

27 impaired airflow which is reversible by treatment with aerosolized albuterol; increased 

28 abundance of immune cells, especially eosinophils, in airway exudates and bronchial lavage; and 

29 development of nonspecific airway responsiveness (NHLBI, 2007). These studies and others 

30 have demonstrated changes in pulmonary function and airway morphology in adult and infant 

31 nonhuman primates repeatedly exposed to environmentally relevant concentrations of O3 (ISA, 

32 section 7.2.3.2, 2013). 

33 The initial observations in adult nonhuman primates have been expanded in a series of 

34 experiments using infant rhesus monkeys repeatedly exposed to 0.5 ppm O3 starting at 1 month 
3-45

1 of age (Plopper et al., 2007). The purpose of these studies was to determine if a cyclic regimen of 

2 O3 inhalation would amplify the allergic responses and structural remodeling associated with 

3 allergic sensitization and inhalation in the infant rhesus monkey. After several episodic 

4 exposures of infant monkeys to O3, they observed a significant increase in the baseline airway 

5 resistance, which was accompanied by a small increase in airway responsiveness to inhaled 

6 histamine (Schelegle et al., 2003), although neither measurement was statistically different from 

7 filtered air control values. Exposure of animals to inhaled house dust mite antigen alone also 

8 produced small but not statistically significant changes in baseline airway resistance and airway 

9 responsiveness, whereas the combined exposure to both (O3 + antigen) produced statistically 

10 significant and greater than additive changes in both functional measurements. This nonhuman 

11 primate evidence of an O3-induced change in airway resistance and responsiveness provides 

12 biological plausibility of long-term exposure, or repeated short-term exposures, to O3

13 contributing to the effects of asthma in children. 

14 To understand which conducting airways and inflammatory mechanisms are involved in 

15 O3-induced airway hyperresponsiveness in the infant rhesus monkey, results of a follow-up study 

16 (Joad et al., 2006) suggest that effect of O3 on airway responsiveness occurs predominantly in 

17 the smaller bronchioles, where dosimetric models indicate the dose would be higher. 

18 The functional changes in the conducting airways were accompanied by a number of cellular and 

19 morphological changes, including a significant 4-fold increase in eosinophils. Thus, these studies 

20 demonstrate both functional and cellular changes in the lung of infant monkeys after cyclic 

21 exposure to 0.5 ppm O3, providing relevant information to understanding the potentially 

22 damaging effects of ambient O3 exposure on the respiratory tract of children. 

23 In addition, significant structural changes in the respiratory tract development, during 

24 which conducting airways increase in diameter and length, have been observed in infant rhesus 

25 monkeys after cyclic exposure to O3 (Fanucchi et al., 2006). Observed changes included more 

26 proximal first alveolar outpocketing, decreases in the diameter and length of the terminal and 

27 respiratory bronchioles, increases in mucus-producing goblet cell mass, alterations in smooth 

28 muscle orientation in the respiratory bronchioles, epithelial nerve fiber distribution, and 

29 basement membrane zone morphometry. The latter effects are noteworthy because of their 

30 potential contribution to airway obstruction and airway hyperresponsiveness which are central 

31 features of asthma. A number of studies in both non-human primates and rodents demonstrate 

32 that O3 exposure can increase collagen synthesis and deposition, including fibrotic-like changes 

33 in the lung (ISA, section 7.2.3.2, U.S. EPA, 2013). 

34 Collectively, evidence from animal studies strongly suggests that chronic O3 exposure is 

35 capable of damaging the distal airways and proximal alveoli, resulting in lung tissue remodeling 
3-46

1 and leading to apparent irreversible changes. Potentially, persistent inflammation and interstitial 

2 remodeling play an important role in the progression and development of chronic lung disease. 

3 Further discussion of the modes of action that lead to O3-induced morphological changes can be 

4 found in Section 5.3.7 of the ISA (U.S. EPA, 2013). Discussion of mechanisms involved in 

5 lifestage susceptibility and developmental effects can be found in Section 5.4.2.4 of the ISA 

6 (U.S. EPA, 2013). The findings reported in chronic animal studies offer insight into potential 

7 biological mechanisms for the suggested association between seasonal O3 exposure and reduced 

8 lung function development in children as observed in epidemiologic studies (see Section 7.2.3.1). 

9 Respiratory Mortality 

10 A limited number of epidemiologic studies have assessed the relationship between long11 term exposure to O3 and mortality in adults. The 2006 O3 AQCD concluded that an insufficient 

12 amount of evidence existed “to suggest a causal relationship between chronic O3 exposure and 

13 increased risk for mortality in humans” (U.S. EPA, 2006b). Though total and cardio-pulmonary 

14 mortality were considered in these studies, respiratory mortality was not specifically considered. 

15 In the most recent follow-up analysis of the ACS cohort (Jerrett et al., 2009), cardiopulmonary 

16 deaths were separately subdivided into respiratory and cardiovascular deaths, rather than 

17 combined as in the Pope et al. (2002) work. Increased O3 exposure was associated with the risk 

18 of death from respiratory causes, and this effect was robust to the inclusion of PM2.5. The 

19 association between increased O3 concentrations and increased risk of death from respiratory 

20 causes was insensitive to the use of different models and to adjustment for several ecologic 

21 variables considered individually. Additionally, a recent multi-city time series study (Zanobetti 

22 and Schwartz, 2011), which followed (from 1985 to 2006) four cohorts of Medicare enrollees 

23 with chronic conditions that might predispose to O3-related effects, observed an association 

24 between long-term (warm season) exposure to O3 and elevated risk of mortality in the cohort that 

25 had previously experienced an emergency hospital admission due to COPD. A key limitation of 

26 this study is the inability to control for PM2.5, because data were not available in these cities until 

27 1999. 

28 3.1.2.3 Total Mortality – Short-term Exposures 

29 To what extent does the currently available scientific evidence, including related 

30 uncertainties, strengthen or alter our understanding from the last review of 

31 mortality attributable to short-term O3 exposures? 

32 The 2006 O3 AQCD concluded that the overall body of evidence was highly suggestive 

33 that short-term exposure to O3 directly or indirectly contributes to nonaccidental and 

34 cardiopulmonary-related mortality in adults, but additional research was needed to more fully 
3-47

1 establish underlying mechanisms by which such effects occur (U.S. EPA, 2013, p. 2-18). In 

2 building on the 2006 evidence, the ISA states the following (U.S. EPA, 2013, p. 6-261). 

3 The evaluation of new multicity studies that examined the association between 

4 short-term O3 exposures and mortality found evidence that supports the 

5 conclusions of the 2006 AQCD. These new studies reported consistent positive 

6 associations between short-term O3 exposure and all-cause (non-accidental) 

7 mortality, with associations persisting or increasing in magnitude during the 

8 warm season, and provide additional support for associations between O3

9 exposure and cardiovascular and respiratory mortality 

10 The 2006 O3 AQCD reviewed a large number of time-series studies of associations 

11 between short-term O3 exposures and total mortality including single- and multicity studies, and 

12 meta-analyses. In the large U.S. multicity studies that examined all-year data, effect estimates 

13 corresponding to single-day lags ranged from a 0.5-1% increase in all-cause (nonaccidental) total 

14 mortality per a 20 ppb (24-hour), 30 ppb (8-hour maximum), or 40 ppb (1-hour maximum) 

15 increase in ambient O3 (U.S. EPA, 2013, section 6.6.2). Available studies reported some 

16 evidence for heterogeneity in O3 mortality risk estimates across cities and across studies. Studies 

17 that conducted seasonal analyses reported larger O3 mortality risk estimates during the warm 

18 season. Overall, the 2006 O3 AQCD identified robust associations between various measures of 

19 daily ambient O3 concentrations and all-cause mortality, which could not be readily explained by 

20 confounding due to time, weather, or copollutants. With regard to cause-specific mortality, 

21 consistent positive associations were reported between short-term O3 exposure and 

22 cardiovascular mortality, with less consistent evidence for associations with respiratory 

23 mortality. The majority of the evidence for associations between O3 and cause-specific mortality 

24 were from single-city studies, which had small daily mortality counts and subsequently limited 

25 statistical power to detect associations. The 2006 O3 AQCD concluded that “the overall body of 

26 evidence is highly suggestive that O3 directly or indirectly contributes to non-accidental and 

27 cardiopulmonary-related mortality” (U.S. EPA, 2012a, section 6.6.1). 

28 Recent studies have strengthened the body of evidence that supports the association 

29 between short-term O3 concentrations and mortality in adults. This evidence includes a number 

30 of studies reporting associations with non-accidental as well as cause-specific mortality. Multi31 continent and multicity studies have consistently reported positive and statistically significant 

32 associations between short-term O3 concentrations and all-cause mortality, with evidence for 

33 larger mortality risk estimates during the warm or summer months (Figure 3-3 below, reprinted 

34 from the ISA) (U.S. EPA, 2013, Figure 6-27; Table 6-42). Similarly, evaluations of cause-
3-48

1 specific 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).24 2 

 

24Respiratory mortality is discussed in more detail above. 
December 2013 3-49 Draft - Do Not Quote or Cite

1

2

3

4 

Figure 3-3. Summary of mortality risk estimates for short-term O3 and all-cause (nonaccidental) mortality.25 5

 

25Reprinted 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

0-1

0-1

0

0-6

0-6

DL(0-2 )

0-1

DL(0-2 )

DL(0-6 )

DL(0-2 )

DL(0-2 )

0-1

0-6

0

0

0-3

0

0-1

0-2

DL(0-2 )

DL(0-2 )

DL(0-2 )

DL(0-2 )

DL(0-5 ) ►

All-Year

Summer

-1 1 3 5 7 9 1 1

% Increase
3-50

1 In assessing the evidence for O3-related mortality, the 2006 AQCD also noted that 

2 multiple uncertainties remained regarding the relationship between short-term O3 concentrations 

3 and mortality, including the extent of residual confounding by co-pollutants; characterization of 

4 the factors that modify the O3-mortality association; the appropriate lag structure for identifying 

5 O3-mortality effects; and the shape of the O3-mortality concentration-response function and 

6 whether a threshold exists. Many of the studies, published since the last review, have attempted 

7 to address one or more of these uncertainties. The ISA (U.S. EPA, 2013; Section 6.6.2) discusses 

8 the extent to which recent studies have evaluated these uncertainties in the relationship between 

9 O3 and mortality. 

10 In particular, recent studies have evaluated different statistical approaches to examine the 

11 shape of the O3-mortality concentration-response relationship and to evaluate whether a 

12 threshold exists for O3-related mortality. In an analysis of the NMMAPS data, Bell et al. (2006) 

13 evaluated the potential for a threshold in the O3-mortality relationship. The authors reported 

14 positive and statistically significant associations with mortality in a variety of restricted analyses, 

15 including analyses restricted to days with 24-hour area-wide average O3 concentrations below 

16 60, 55, 50, 45, 40, 35, and 30 ppb. In these restricted analyses O3 effect estimates were of similar 

17 magnitude, were statistically significant, and had similar statistical precision. In analyses 

18 restricted to days with 24-hour average O3 concentrations below 25 ppb, the O3 effect estimate 

19 was similar in magnitude to the effect estimates resulting from analyses with the higher cutoffs, 

20 but had somewhat lower statistical precision, with the estimate approaching statistical 

21 significance (i.e., based on observation of Figure 2 in Bell et al., 2006). In analyses restricted to 

22 days with lower 24-hour average O3 concentrations (i.e., below 20 and 15 ppb), effect estimates 

23 were similar in magnitude to analyses with higher cutoffs, but with notably less statistical 

24 precision, and were not statistically significant (i.e., confidence intervals included no O3-

25 associated mortality based on observation of Figure 2 in Bell et al., 2006). Ozone was no longer 

26 positively associated with mortality when the analysis was restricted to days with 24-hour O3

27 concentrations below 10 ppb. Given the relatively small number of days included in these 

restricted analyses, especially for cut points of 20 ppb and below,26 28 statistical uncertainty is 

29 increased. 

30 Bell et al. (2006) also evaluated the shape of the concentration-response relationship 

31 between O3 and mortality. Although the results of this analysis suggested the lack of threshold in 

32 the O3-mortality relationship, the ISA noted that it is difficult to interpret such a curve because: 

33 (1) there is uncertainty around the shape of the concentration-response curve at 24-hour average 

 

26For example, Bell et al. (2006) reported that for analyses restricted to 24-hour O3 concentrations at or below 20 

ppb, 73% of days were excluded on average across the 98 communities. 
3-51

1 O3 concentrations generally below 20 ppb and (2) the concentration-response curve does not take 

2 into consideration the heterogeneity in O3-mortality risk estimates across cities (U.S. EPA, 2013, 

3 section 6.6.2.3). 

4 Several additional studies have used the NMMAPS dataset to evaluate the concentration5 response relationship between short-term O3 concentrations and mortality. For example, using 

6 the same data as Bell et al. (2006), Smith et al. (2009) conducted a subset analysis, but instead of 

7 restricting the analysis to days with O3 concentrations below a cutoff the authors only included 

8 days above a defined cutoff. The results of this analysis were consistent with those reported by 

9 Bell et al. (2006). Specifically, the authors reported consistent positive associations for all cutoff 

10 concentrations up to concentrations where the total number of days available were so limited that 

11 the variability around the central estimate was increased (U.S. EPA, 2013, section 6.6.2.3). In 

12 addition, using NMMAPS data for 1987-1994 for Chicago, Pittsburgh, and El Paso, Xia and 

13 Tong (2006) reported evidence for a threshold around a 24-hour average O3 concentration of 

14 25 ppb, though the threshold values estimated in the analysis were sometimes in the range of 

15 where data density was low (U.S. EPA, 2013, section 6.6.2.3). Stylianou and Nicolich (2009) 

16 examined the existence of thresholds following an approach similar to Xia and Tong (2006) 

17 using data from NMMAPS for nine major U.S. cities (i.e., Baltimore, Chicago, Dallas/Fort 

18 Worth, Los Angeles, Miami, New York, Philadelphia, Pittsburgh, and Seattle) for the years 

19 1987-2000. The authors reported that the estimated O3-mortality risks varied across the nine 

20 cities, with the models exhibiting apparent thresholds in the 10-45 ppb range for O3 (24-hour 

21 average). However, given the city-to-city variation in risk estimates, combining the city-specific 

22 estimates into an overall estimate complicated the interpretation of the results. Additional studies 

23 in Europe, Canada, and Asia did not report the existence of a threshold (Katsouyanni et al., 

24 2009), with inconsistent and/or inconclusive results across cities, or a non-linear relationship in 

25 the O3-mortality concentration-response curve (Wong et al., 2010). 

26 3.1.2.4 Cardiovascular effects – Short-term Exposure 

27 To what extent does the currently available scientific evidence, including related 

28 uncertainties, strengthen or alter our understanding from the last review of 

29 cardiovascular effects attributable to short-term O3 exposures? 

30 A relatively small number of studies have examined the potential effect of short-term O3

31 exposure on the cardiovascular system. The 2006 O3 AQCD (U.S. EPA, 2006, p.8-77) concluded 

32 that “O3 directly and/or indirectly contributes to cardiovascular-related morbidity” but added that 

33 the body of evidence was limited. This conclusion was based on a controlled human exposure 

34 study that included hypertensive adult males; a few epidemiologic studies of physiologic effects, 
3-52

1 heart rate variability, arrhythmias, myocardial infarctions, and hospital admissions; and 

2 toxicological studies of heart rate, heart rhythm, and blood pressure. 

3 More recently, the body of scientific evidence available that has examined the effect of 

4 O3 on the cardiovascular system has expanded. There is an emerging body of animal 

5 toxicological evidence demonstrating that short-term exposure to O3 can lead to autonomic 

6 nervous system alterations (in heart rate and/or heart rate variability) and suggesting that 

7 proinflammatory signals may mediate cardiovascular effects. Interactions of O3 with respiratory 

8 tract components result in secondary oxidation product formation and subsequent production of 

9 inflammatory mediators, which have the potential to penetrate the epithelial barrier and to initiate 

10 toxic effects systemically. In addition, animal toxicological studies of long-term exposure to O3

11 provide evidence of enhanced atherosclerosis and ischemia/reperfusion (I/R) injury, 

12 corresponding with development of a systemic oxidative, proinflammatory environment. Recent 

13 experimental and epidemiologic studies have investigated O3-related cardiovascular events and 

14 are summarized in Section 6.3 of the ISA (U.S. EPA, 2013, Section 6.3). Overall, the ISA 

15 summarized the evidence in this review as follows (U.S. EPA, 2013, p. 6-211). 

16 In conclusion, animal toxicological studies demonstrate O3-induced 

17 cardiovascular effects, and support the strong body of epidemiologic evidence 

18 indicating O3-induced cardiovascular mortality. Animal toxicological and 

19 controlled human exposure studies provide evidence for biologically plausible 

20 mechanisms underlying these O3-induced cardiovascular effects. However, a lack 

21 of coherence with epidemiologic studies of cardiovascular morbidity remains an 

22 important uncertainty.

23 Animal toxicological studies support that short-term O3 exposure can lead to 

24 cardiovascular morbidity. Animal studies provide evidence for both increased and decreased 

25 heart rate (HR), however it is uncertain if O3-induced reductions in HR are relevant to humans. 

26 Animal studies also provide evidence for increased heart rate variability (HRV), arrhythmias, 

27 vascular disease and injury following short-term O3 exposure. In addition, a series of studies 

28 highlight the role of genetic variability and age in the induction of effects and attenuation of 

29 responses to O3 exposure. 

30 Biologically plausible mechanisms have been described for the cardiovascular effects 

31 observed in animal exposure studies (U.S. EPA, 2013, Section 5.3.8). Evidence that 

32 parasympathetic pathways may underlie cardiac effects is described in more detail in Section 

33 5.3.2 of the ISA (U.S. EPA, 2013). Recent studies suggest that O3 exposure may disrupt the 

34 endothelin system that constricts blood vessels and increase blood pressure, which can result in 

35 an increase in HR, HRV; and disrupt the NO ̇system and the production of atrial natriuretic 

36 factor (ANF), vasodilators that reduce blood pressure. Additionally, O3 may increase oxidative 

37 stress and vascular inflammation promoting the progression of atherosclerosis and leading to 
3-53

1 increased susceptibility to I/R injury. As O3 reacts quickly with the ELF and does not translocate 

2 to the heart and large vessels, studies suggest that the cardiovascular effects exhibited could be 

3 caused by secondary oxidation products resulting from O3 exposure. However, direct evidence of 

4 translocation of O3 reaction products to the cardiovascular system has not been demonstrated in 

5 vivo. Alternatively, extrapulmonary release of diffusible mediators (such as cytokines or 

6 endothelins) may initiate or propagate inflammatory responses throughout the body leading to 

7 the cardiovascular effects reported in toxicology studies. Ozone reacts within the lung to induce 

8 pulmonary inflammation and the influx and activation of inflammatory cells, resulting in a 

9 cascading proinflammatory state, and may lead to the extrapulmonary release of diffusible 

10 mediators that could result in cardiovascular injury. 

11 Controlled human exposures studies discussed in previous AQCDs have not 

12 demonstrated any consistent extrapulmonary effects. In this review, evidence from controlled 

13 human exposure studies suggests cardiovascular effects in response to short-term O3 exposure 

14 (see ISA, U.S. EPA, 2013, Section 6.3.1) and provides some coherence with evidence from 

15 animal toxicology studies. Controlled human exposure studies also support the animal 

16 toxicological studies by demonstrating O3-induced effects on blood biomarkers of systemic 

17 inflammation and oxidative stress, as well as changes in biomarkers that can indicate a 

18 prothrombogenic response to O3. Increases and decreases in high frequency HRV have been 

19 reported following relatively low (120 ppb during rest) and high (300 ppb with exercise) O3

20 exposures, respectively. These changes in cardiac function observed in animal and human studies 

21 provide preliminary evidence for O3-induced modulation of the autonomic nervous system 

22 through the activation of neural reflexes in the lung (see ISA,U.S. EPA 2013, Section 5.3.2). 

23 Overall, the ISA concludes that the available body of epidemiologic evidence examining 

24 the relationship between short-term exposures to O3 concentrations and cardiovascular morbidity 

25 is inconsistent (U.S. EPA, 2013, Section 6.3.2.9). Across studies, different definitions, (i.e., ICD26 9 diagnostic codes) were used for both all-cause and cause-specific cardiovascular morbidity 

27 (ISA, U.S. EPA, 2013, see Tables 6-35 to 6-39), which may contribute to inconsistency in 

28 results. However, within diagnostic categories, no consistent pattern of association was found 

29 with O3. Generally, the epidemiologic studies used nearest air monitors to assess O3

30 concentrations, with a few exceptions that used modeling or personal exposure monitors. 

31 The inconsistencies in the associations observed between short-term O3 and cardiovascular 

32 disease (CVD) morbidities are unlikely to be explained by the different exposure assignment 

33 methods used (see Section 4.6, ISA, U.S. EPA 2013). The wide variety of biomarkers considered 

34 and the lack of consistency among definitions used for specific cardiovascular disease endpoints 

35 (e.g., arrhythmias, HRV) make comparisons across studies difficult. 
3-54

1 Despite the inconsistent evidence for an association between O3 concentration and CVD 

2 morbidity, mortality studies indicate a consistent positive association between short-term O3

3 exposure and cardiovascular mortality in multicity studies and in a multicontinent study. When 

4 examining mortality due to cardiovascular disease, epidemiologic studies consistently observe 

5 positive associations with short-term exposure to O3. Additionally, there is some evidence for an 

6 association between long-term exposure to O3 and mortality, although the association between 

7 long-term ambient O3 concentrations and cardiovascular mortality can be confounded by other 

8 pollutants as evident by a study of cardiovascular mortality that reported no association after 

9 adjustment for PM2.5 concentrations. The ISA (U.S. EPA 2013, section 6.3.4) states that taken 

10 together, the overall body of evidence across the animal and human studies is sufficient to 

11 conclude that there is likely to be a causal reationship between relevant short-term exposures to 

12 O3 and cardiovascular system effects. 

13 3.1.3 Adversity of Effects 

14 In this section we address the following question: 

15 To what extent does the currently available scientific evidence expand our 

16 understanding of the adversity of O3-related health effects?

17 In making judgments as to when various O3-related effects become regarded as adverse 

18 to the health of individuals, in previous NAAQS reviews staff has relied upon the guidelines 

19 published by the American Thoracic Society (ATS) and the advice of CASAC. In 2000, the ATS 

20 published an official statement on “What Constitutes an Adverse Health Effect of Air 

21 Pollution?” (ATS, 2000), which updated and built upon its earlier guidance (ATS, 1985). The 

22 earlier guidance defined adverse respiratory health effects as “medically significant physiologic 

23 changes generally evidenced by one or more of the following: (1) interference with the normal 

24 activity of the affected person or persons, (2) episodic respiratory illness, (3) incapacitating 

25 illness, (4) permanent respiratory injury, and/or (5) progressive respiratory dysfunction”, while 

26 recognizing that perceptions of “medical significance” and “normal activity” may differ among 

27 physicians, lung physiologists and experimental subjects (ATS, 1985). The 2000 ATS guidance 

28 builds upon and expands the 1985 definition of adversity in several ways. The guidance 

29 concludes that transient, reversible loss of lung function in combination with respiratory 

30 symptoms should be considered adverse. There is also a more specific consideration of 

31 population risk (ATS, 2000). Exposure to air pollution that increases the risk of an adverse effect 

32 to the entire population is adverse, even though it may not increase the risk of any individual to 

33 an unacceptable level. For example, a population of asthmatics could have a distribution of lung 

34 function such that no individual has a level associated with significant impairment. Exposure to 

35 air pollution could shift the distribution to lower levels that still do not bring any individual to a 
3-55

1 level that is associated with clinically relevant effects. However, this would be considered to be 

2 adverse because individuals within the population would have diminished reserve function, and 

3 therefore would be at increased risk to further environmental insult (U.S. EPA, 2013, p. lxxi; and 

4 SO2 NAAQS review, 75 FR at 35526/2, June 22, 2010).

5 The ATS also concluded that elevations of biomarkers such as cell types, cytokines and 

6 reactive oxygen species may signal risk for ongoing injury and more serious effects or may 

7 simply represent transient responses, illustrating the lack of clear boundaries that separate 

8 adverse from nonadverse events. More subtle health outcomes also may be connected 

9 mechanistically to health effects that are clearly adverse, so that small changes in physiological 

10 measures may not appear clearly adverse when considered alone, but may be part of a coherent 

11 and biologically plausible chain of related health outcomes that include responses that are clearly 

12 adverse, such as mortality (section 3.1.2.1, above). 

13 In this review, the new evidence provides further support for relationships between O3

14 exposures and a spectrum of health effects, including effects that meet the ATS criteria for being 

15 adverse (ATS, 1985 and 2000). The ISA judgment that there is a causal relationship between 

16 short-term O3 exposure and a full range of respiratory effects, including respiratory morbidity 

17 (e.g., lung function decrements, respiratory symptoms, inflammation, hospital admissions, and 

18 emergency department visits) and mortality, provides support for concluding that short-term O3

19 exposure is associated with adverse effects (U.S. EPA 2013, section 2.5.2). Overall, including 

20 new evidence of cardiovascular system effects, the evidence supporting an association between 

21 short-term O3 exposures and total (non-accidental, cardiopulmonary) respiratory mortality is 

22 stronger in this review (U.S. EPA 2013, section 2.5.2). And the judgment of likely causal 

23 associations between long-term measures of O3 exposure and respiratory effects such as new24 onset asthma, prevalence of asthma, asthma symptoms and control, and asthma hospital 

25 admissions provides support for concluding that long-term O3 exposure is associated with 

26 adverse effects ranging from episodic respiratory illness to permanent respiratory injury or 

27 progressive respiratory decline (U.S. EPA 2013, section 7.2.8).

28 This review provides additional evidence of O3-attributable effects that are clearly 

29 adverse, including premature mortality. Application of the ATS guidelines to the least serious 

30 category of effects related to ambient O3 exposures, which are also the most numerous and 

31 therefore are also important from a public health perspective, involves judgments about which 

32 medical experts on CASAC panels and public commenters have in the past expressed diverse 

33 views. To help frame such judgments, EPA staff defined gradations of individual functional 

34 responses (e.g., decrements in FEV1 and airway responsiveness) and symptomatic responses 

35 (e.g., cough, chest pain, wheeze), together with judgments as to the potential impact on 

36 individuals experiencing varying degrees of severity of these responses. These gradations were 
3-56

1 used in the 1997 O3 NAAQS review and slightly revised in the 2008 review (U.S. EPA, 1996, p. 

2 59; 2007, p.3-72; 72 FR 37849, July 11, 2007). These gradations and impacts are summarized in 

3 Tables 3-2 and 3-3 in the 2007 O3 Staff Paper (U.S. EPA, 2007, pp.3-74 to 3-75). 

4 For active healthy people, including children, moderate levels of functional responses 

5 (e.g., FEV1 decrements of ≥ 10% but < 20%, lasting 4 to 24 hours) and/or moderate symptomatic 

6 responses (e.g., frequent spontaneous cough, marked discomfort on exercise or deep breath, 

7 lasting 4 to 24 hours) would likely interfere with normal activity for relatively few sensitive 

8 individuals (U.S. EPA, 2007, p.3-72; 72 FR 37849, July 11, 2007); whereas large functional 

9 responses (e.g., FEV1 decrements ≥ 20%, lasting longer than 24 hours) and/or severe 

10 symptomatic responses (e.g., persistent uncontrollable cough, severe discomfort on exercise or 

11 deep breath, lasting longer than 24 hours) would likely interfere with normal activities for many 

12 sensitive individuals (U.S. EPA, 2007, p.3-72; 72 FR 37849, July 11, 2007) and therefore would 

13 be considered adverse under ATS guidelines. For the purpose of estimating potentially adverse 

14 lung function decrements in active healthy people in the 2008 O3 NAAQS review, the CASAC 

15 panel for that review indicated that a focus on the mid to upper end of the range of moderate 

16 levels of functional responses is most appropriate (e.g., FEV1 decrements ≥ 15% but < 20%) 

17 (Henderson, 2006; 2007 Staff Paper, p. 3-76). However, for children and adults with lung 

18 disease, even moderate functional (e.g., FEV1 decrements ≥ 10% but < 20%, lasting up to 24 

19 hours) or symptomatic responses (e.g., frequent spontaneous cough, marked discomfort on 

20 exercise or with deep breath, wheeze accompanied by shortness of breath, lasting up to 24 hours) 

21 would likely interfere with normal activity for many individuals, and would likely result in 

22 additional and more frequent use of medication (U.S. EPA, 2007, p.3-72; 72 FR 37849, July 11, 

23 2007). For people with lung disease, large functional responses (e.g., FEV1 decrements ≥ 20%, 

24 lasting longer than 24 hours) and/or severe symptomatic responses (e.g., persistent 

25 uncontrollable cough, severe discomfort on exercise or deep breath, persistent wheeze 

26 accompanied by shortness of breath, lasting longer than 24 hours) would likely interfere with 

27 normal activity for most individuals and would increase the likelihood that these individuals 

28 would seek medical treatment (U.S. EPA, 2007, p.3-72; 72 FR 37849, July 11, 2007). In the last 

29 O3 NAAQS review, for the purpose of estimating potentially adverse lung function decrements 

30 in people with lung disease the CASAC panel indicated that a focus on the lower end of the 

31 range of moderate levels of functional responses is most appropriate (e.g., FEV1 decrements 

32 ≥10%) (Henderson, 2006; 2007 Staff Paper, p. 3-76). In addition, in the reconsideration of the 

33 2008 final decision, CASAC stated that “[a] 10% decrement in FEV1 can lead to respiratory 

34 symptoms, especially in individuals with pre-existing pulmonary or cardiac disease. For 

35 example, people with chronic obstructive pulmonary disease have decreased ventilatory reserve 
3-57

1 (i.e., decreased baseline FEV1) such that a ≥10% decrement could lead to moderate to severe 

2 respiratory symptoms” (Samet, 2011) (section 3.1.2.1, above). 

3 In judging the extent to which these impacts represent effects that should be regarded as 

4 adverse to the health status of individuals, in previous NAAQS reviews we also considered 

5 whether effects were experienced repeatedly during the course of a year or only on a single 

6 occasion (Staff Paper, U.S. EPA, 2007). Although some experts would judge single occurrences 

7 of moderate responses to be a ‘‘nuisance,’’ especially for healthy individuals, a more general 

8 consensus view of the adversity of such moderate responses emerges as the frequency of 

9 occurrence increases. Thus it has been judged that repeated occurrences of moderate responses, 

10 even in otherwise healthy individuals, may be considered to be adverse since they could well set 

11 the stage for more serious illness (61 FR 65723). The CASAC panel in the 1997 NAAQS review 

12 expressed a consensus view that these “criteria for the determination of an adverse physiological 

13 response were reasonable” (Wolff, 1995). In the review completed in 2008, estimates of repeated 

14 occurrences continued to be an important public health policy factor in judging the adversity of 

15 moderate lung function decrements in healthy and asthmatic people (72 FR 37850, July 11, 

16 2007). 

17 Evidence new to this review indicates that 6.6-hour exposures to 60 ppb O3 during 

18 moderate exertion can result in pulmonary inflammation in healthy adults. As discussed in 

19 section 3.1.2 above, the initiation of inflammation can be considered as evidence that injury has 

20 occurred. Inflammation induced by a single O3 exposure can resolve entirely, but continued 

21 acute inflammation can evolve into a chronic inflammatory state (ISA, U.S. EPA, 2013, p. 6-76), 

22 which is clearly adverse. Therefore, like moderate lung function decrements, whether 

23 inflammation is experienced repeatedly during the course of a year or only on a single occasion 

24 is judged by staff to be reasonable criteria for determining adverse inflammatory effects 

25 attributable to O3 exposures at 60 ppb. 

26 Responses measured in controlled human exposure studies indicate that the range of 

27 effects elicited in humans exposed to ambient O3 concentrations include: decreased inspiratory 

28 capacity; mild bronchoconstriction; rapid, shallow breathing pattern during exercise; and 

29 symptoms of cough and pain on deep inspiration (EPA, 2013, section 6.2.1.1). Some young, 

30 healthy adults exposed to O3 concentrations ≥ 60 ppb, while engaged in 6.6 hours of intermittent 

31 moderate exertion, develop statistically significant reversible, transient decrements in lung 

32 function, symptoms of breathing discomfort, and inflammation if minute ventilation or duration 

33 of exposure is increased sufficiently (EPA, 2013, section 6.2.1.1). Among healthy subjects there 

34 is considerable interindividual variability in the magnitude of the FEV1 responses, but averaged 

35 across studies at 60 ppb (EPA, 2013, pp. 6-17 to 6-18), 10% of healthy subjects had >10% FEV1

36 decrements. The combination of lung function decrements and respiratory symptoms, which has 
3-58

1 been considered adverse in previous reviews, has been demonstrated in healthy adults following 

2 prolonged (6.6 hour) exposures, while at intermittent moderate exertion, to 70 ppb. For these 

3 types of effects, information from controlled human exposure studies, which provides an 

4 indication of the magnitude and thus adversity of effects at different O3 concentrations, 

5 combined with estimates of occurrences in the population from the HREA, provide information 

6 about their importance from a policy perspective. 

7 3.1.4 Ozone Concentrations Associated With Health Effects 

8 In evaluating O3 exposure concentrations reported to result in health effects, within the 

9 context of the adequacy of the current standard, we first consider the following specific question: 

10 To what extent does the currently available scientific evidence indicate morbidity 

11 and/or mortality attributable to exposures to O3 concentrations lower than 

12 previously reported or that would meet the current standard?

13 In addressing this question, we characterize the extent to which O3-attributable effects have been 

14 reported over the ranges of O3 exposure concentrations evaluated in controlled human exposure 

15 studies and over the distributions of ambient O3 concentrations in locations where epidemiologic 

16 studies have been conducted.

17 3.1.4.1 Concentrations in Controlled Human Exposure Studies and in Epidemiologic 

18 Panel Studies 

19 In considering what the currently available evidence indicates with regard to effects 

20 associated with exposure concentrations lower than those identified in the last review, or that 

21 could meet the current standard, we first consider the evidence from controlled human exposure 

22 studies and epidemiologic panel studies. This evidence is assessed in section 6.2 of the ISA and 

23 is summarized in section 3.1.2 above. Controlled human exposure studies have generally been 

24 conducted with young, healthy adults, and have evaluated exposure durations less than 8 hours. 

25 Epidemiologic panel studies have evaluated a wider range of study populations, including 

26 children, and have generally evaluated associations with O3 concentrations averaged over several 

hours (U.S. EPA, 2013, section 6.2.1.2).27 27

28 As summarized above (section 3.1.2.1), and as discussed in detail in the ISA (U.S. EPA, 

29 2013, section 6.2), a large number of controlled human exposure studies have reported lung 

30 function decrements, respiratory symptoms, airway inflammation, airway hyperresponsiveness, 

31 and/or impaired lung host defense in young, healthy adults engaged in moderate, intermittent 

 

27In 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). 
3-59

1 exertion, following 6.6-hour O3 exposures. These studies have consistently reported such effects 

2 following exposures to O3 concentrations of 80 ppb or greater. Available studies have also 

3 evaluated some of these effects (i.e., lung function decrements, respiratory symptoms, airway 

4 inflammation) following exposures to O3 concentrations below 75 ppb. Table 3-1 highlights the 

5 group mean results of individual controlled human exposure studies that have evaluated 

6 exposures of healthy adults to O3 concentrations below 75 ppb. 

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

8 exposures to ozone concentrations below 75 ppb in young, healthy adults.

Endpoint O3 Exposure 

Concentration Study

Statistically

Significant O3-

Induced 

Effect28

FEV1 decrements

70 ppb Schelegle et al., 2009 yes

60 ppb

Kim et al., 2011 yes

Schelegle et al., 2009 no

Adams, 2006 yes29

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 no30

40 ppb Adams, 2006 no

Adams, 2002 no

Airway 

Inflammation 

(neutrophil influx)

60 ppb Kim et al., 2011 yes

9

10 In further evaluating O3-induced FEV1 decrements following exposures to O3

11 concentrations below 75 ppb, the ISA also combined the individual data from multiple studies of 

12 healthy adults exposed for 6.6 hours to 60 ppb O3 (Kim et al., 2011; Brown et al., 2008; Adams, 

13 2006a, 2002, 1998). Based on these data, the ISA reports that 10% of exposed subjects 

14 experienced FEV1 decrements of 10% or more (i.e., abnormal and large enough to be potentially 

 

28Based on study population means. 

29 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). 

30Adams (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. 
3-60

1 adverse for people with pulmonary disease, based on past CASAC advice (section 3.1.3, 

above))31 2 (U.S. EPA, 2013, section 6.2.1.1). Consistent with these findings, recently developed 

3 empirical models predict that the onset of O3-induced FEV1 decrements in healthy adults occurs 

4 following exposures to 60 ppb O3 for 4 to 5 hours while at moderate, intermittent exertion 

5 (Schelegle et al., 2012), and that 9% of healthy adults exposed to 60 ppb O3 for 6.6 hours would 

6 experience FEV1 decrements greater than or equal to 10% (McDonnell et al., 2012) (U.S. EPA, 

7 2013, section 6.2.1.1). When the evidence for O3-induced lung function decrements was taken 

8 together, the ISA concluded that (1) “mean FEV1 is clearly decreased by 6.6-h exposures to 60 

9 ppb O3 and higher concentrations in subjects performing moderate exercise” (U.S. EPA, 2013, p. 

10 6-9) and (2) although group mean decrements following exposures to 60 ppb O3 are biologically 

11 small, “a considerable fraction of exposed individuals experience clinically meaningful 

12 decrements in lung function” (U.S. EPA, 2013, p. 6-20). 

13 In considering the specific question above, we note that controlled human exposure 

14 studies have reported decreased lung function, increased airway inflammation, and increased

15 respiratory symptoms in healthy adults following exposures to O3 concentrations below 75 ppb. 

16 Such impairments in respiratory function have the potential to be adverse, based on ATS 

17 guidelines for adversity and based on previous advice from CASAC (section 3.1.3, above). In 

18 addition, if they become serious enough, these respiratory effects could lead to the types of 

19 clearly adverse effects commonly reported in O3 epidemiologic studies (e.g., respiratory 

20 emergency department visits, hospital admissions). Therefore, following exposures to O3

21 concentrations lower than 75 ppb, controlled human exposure studies have reported respiratory 

22 effects that could be adverse in some individuals, particularly if experienced by members of atrisk populations (e.g., asthmatics, children).32 23

24 In further considering effects following exposures to O3 concentrations below 75 ppb, we 

25 also note that the ISA highlights some epidemiologic panel studies for the extent to which 

26 monitored ambient O3 concentrations reflect exposure concentrations in their study populations 

27 (U.S. EPA, 2013, section 6.2.1.2). Specifically, Table 3-2 below includes O3 panel studies that 

28 have evaluated associations with lung function decrements for O3 concentrations at or below 75 

29 ppb, and that measured O3 concentrations with monitors located in the areas where study 

30 subjects were active (e.g., on site at summer camps or in locations where exercise took place) 

 

31CASAC 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) (section 3.1.3, above). 

32These effects were reported in healthy individuals. It is thus 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). 
3-61

1 (U.S. EPA, 2013, section 6.2.1.2 and Table 6-6). Epidemiologic panel studies have evaluated a 

2 wider range of populations and lifestages than controlled human exposure studies of O3

3 concentrations below 75 ppb (e.g., including children). 

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

5 concentrations below 75 ppb. 

6 Although these studies report health effect associations for different averaging times, and it is not 

7 clear the extent to which specific O3 exposure conditions (i.e., concentrations, durations of 

8 exposure, degrees of activity) were responsible for eliciting reported decrements, they are 

9 consistent with the findings of the controlled human exposure studies discussed above. 

10 Specifically, the epidemiologic panel studies in Table 3-2 indicate O3-associated lung function 

11 decrements when on-site monitored concentrations (ranging from minutes to hours) were below 

12 75 ppb, with the evidence becoming less consistent at lower O3 concentrations.

13 3.1.4.2 Concentrations in Epidemiologic Studies – Short-term Metrics 

14 We next consider distributions of ambient O3 concentrations in locations where 

15 epidemiologic studies have evaluated O3-associated hospital admissions, emergency department 

16 visits, and/or mortality. When considering epidemiologic studies within the context of the current 

17 standard, we emphasize those studies conducted in the U.S. and Canada. Such studies reflect air 

18 quality and exposure patterns that are likely more typical of the U.S. population than the air 

19 quality and exposure patterns reflected in studies conducted outside the U.S. and Canada (section 

Study Population O3 Concentrations Statistically Significant Association 

with Lung Function Decrements 

Spektor et al. 

(1988)

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
3-62

1.3.1.2, above). 1

33 We also emphasize studies reporting associations with effects judged in the 

2 ISA to be robust to confounding by other factors, including co-occurring air pollutants. In 

3 addition to these factors, we consider the statistical precision of study results, the extent to which 

4 studies report associations in at-risk populations, and the extent to which the biological 

5 plausibility of associations at various ambient O3 concentrations is supported by controlled 

6 human exposure and/or animal toxicological studies. These considerations help inform the range 

7 of ambient O3 concentrations over which we have the most confidence in O3-associated health 

8 effects, and the range of concentrations over which our confidence in such associations is 

9 appreciably lower. We place particular emphasis on characterizing those portions of distributions 

10 of ambient O3 concentrations likely to meet the current standard. 

11 In our consideration of these issues, we first address the following question: 

12 To what extent have U.S. and Canadian epidemiologic studies reported 

13 associations with mortality or morbidity in locations that would have met the 

14 current O3 standard during the study period? 

15 Addressing this question can provide important insights into the extent to which O3-health effect 

16 associations are present for distributions of ambient O3 concentrations that would be allowed by 

17 the current standard. To the extent O3 health effect associations are reported in study areas that 

18 would have met the current standard, we have greater confidence that the current standard could 

19 allow the clearly adverse O3-associated effects indicated by those studies (e.g., mortality, 

hospital admissions, emergency department visits).34 20

21 We note that epidemiologic studies evaluate statistical associations between variation in 

22 the incidence of health outcomes and variation in ambient O3 concentrations. In many of the O3

23 epidemiologic studies assessed in the ISA, ambient concentrations are averaged across multiple 

24 monitors within study areas, and in some cases over multiple days. These averages are used as 

25 surrogates for the spatial and temporal patterns of O3 exposures in study populations. In this 

26 second draft PA, we refer to these averaged concentrations as “area-wide” O3 concentrations. 

27 We recognize that these epidemiologic studies do not identify the O3 exposures that 

28 population members have experienced, and do not identify the exposures that may be eliciting 

29 the observed health outcomes. Thus, in considering epidemiologic studies of mortality and 

30 morbidity, we are not drawing conclusions regarding single short-duration O3 concentrations in 

31 ambient air that, alone, are eliciting the reported health outcomes. Rather, our focus in this 

32 section is to consider what these studies convey regarding the extent to which health effects may 

 

33Nonetheless, 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. 

34See ATA III, 283 F.3d at 370 (EPA justified in revising NAAQS when health effect associations are observed at 

levels allowed by the NAAQS). 
3-63

1 be occurring (i.e., as indicated by associations) under air quality conditions meeting the current 

2 standard. 

3 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,35 4

5 and mortality (total, respiratory, cardiovascular) from the ISA (studies identified from U.S. EPA, 

6 2013, Table 6-28, section 6.2.8, and Table 6-42) (Appendix 3-D). For each monitor in the areas 

7 evaluated by these studies, we have identified the 3-year averages of the annual 4th highest daily 

maximum 8-hour O3 concentrations (Appendix 3-D).

36 8 To provide perspective on whether study 

9 cities would have met or violated the current O3 NAAQS during the study period, these O3

10 concentrations were compared to the level of the current standard. Based on this approach, a 

11 study city was judged to have met the current standard during the study period if all of the 3-year 

averages of annual 4th 12 highest 8-hour O3 concentrations in that area were below 75 ppb. 

13 Based on these analyses, the large majority of epidemiologic study areas evaluated would 

14 have violated the current standard during study periods (Appendix 3-D). Table 3-3 below 

15 highlights the subset of U.S. and Canadian studies reporting O3 health effect associations in 

16 locations that would have met the current standard during study periods. This includes a U.S. 

17 single-city study that would have met the current standard over the entire study period (Mar and 

18 Koenig, 2009) and four Canadian multicity studies for which the majority of study cities would 

19 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).37 20

21

 

35Given 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. 

36These 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. 

37In 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). 
3-64

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

2 associations in locations that would have met the current standard during 

3 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 mortality38

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

4 As illustrated in Table 3-3, one U.S. single-city study highlighted in the ISA has reported 

5 health effect associations with asthma emergency department visits in a location that would have 

6 met the current standard over the entire study period. In addition, four Canadian multicity studies 

7 reported associations with mortality or morbidity when the majority of study locations would 

8 have met the current standard over the entire study periods. While there is uncertainty in 

9 ascribing the multicity effect estimates reported in these Canadian studies entirely to ambient 

10 concentrations that would have met the current standard (i.e., given that some study locations 

11 would have violated the current standard over at least part of the study period), the information 

12 in Table 3-3 suggests that reported multicity effect estimates are largely influenced by locations 

13 meeting the current standard. Together, these U.S. and Canadian epidemiologic studies suggest a 

14 relatively high degree of confidence in the presence of associations with mortality and morbidity 

15 for ambient O3 concentrations meeting the current standard. 

16 We next consider the extent to which additional epidemiologic studies of mortality or 

17 morbidity (i.e., those conducted in locations that violated the current standard) can also inform 

18 our consideration of adequacy of the current standard. In doing so, we note that health effect 

 

38Katsouyanni et al. (2009) report a positive and statistically significant association with cardiovascular mortality for 

people aged 75 years or older. 
3-65

1 associations reported in epidemiologic studies are influenced by the full distributions of ambient 

2 O3 concentrations, including concentrations below the level of the current standard. We focus on 

3 studies that have explicitly characterized such O3 health effect associations, including confidence 

4 in those associations, for various portions of distributions of ambient O3 concentrations. In doing 

5 so, we consider the following question: 

6 To what extent do analyses from epidemiologic studies indicate confidence in health 

7 effect associations over distributions of ambient O3 concentrations, including at 

8 concentrations lower than previously identified or below the current standard? 

9 We first focus on those studies that have reported confidence intervals around 

10 concentration-response functions over distributions of ambient O3 concentrations. Confidence 

11 intervals around concentration-response functions can provide insights into the range of ambient 

12 concentrations over which the study indicates the most confidence in the reported health effect 

13 associations (i.e., where confidence intervals are narrowest), and into the range of ambient 

14 concentrations below which the study indicates that uncertainty in the nature of such associations 

15 becomes notably greater (i.e., where confidence intervals become markedly wider). The 

16 concentrations below which confidence intervals become markedly wider in such analyses are 

17 intrinsically related to data density, and do not necessarily indicate the absence of an association. 

18 The ISA identifies several epidemiologic studies that have reported confidence intervals 

19 around concentration-response functions in U.S. cities. The ISA concludes that studies generally 

20 indicate a linear concentration-response relationship “across the range of 8-h max and 24 h avg 

21 O3 concentrations most commonly observed in the U.S. during the O3 season” and that “there is 

22 less certainty in the shape of the C-R curve at the lower end of the distribution of O3

23 concentrations” (U.S. EPA, 2013, pp. 2-32 to 2-34). In characterizing the O3 concentrations 

24 below which such certainty decreases, the ISA discusses area-wide O3 concentrations as low as 

25 20 ppb and as high as 40 ppb (U.S. EPA, 2013, section 2.5.4.4). 

26 Consistent with these conclusions, the range of ambient concentrations over which the 

27 evidence indicates the most certainty in concentration-response relationships can vary across 

28 studies. Such variation is likely due at least in part to differences in the O3 metrics evaluated and 

29 differences in the distributions of ambient concentrations and health events. Thus, although 

30 consideration of confidence intervals around concentration-response functions can provide 

31 valuable insights into the ranges of ambient concentrations over which studies indicate the most 

32 confidence in reported health effect associations, there are limitations in the extent to which 

33 these analyses can be generalized across O3 metrics, study locations, study populations, and 

34 health endpoints. 
3-66

1 The ISA emphasizes two U.S. single-city studies that have reported confidence intervals 

2 around concentration-response functions (Silverman and Ito, 2010; Strickland et al., 2010).

3 These studies, and their associated O3 air quality, are discussed below. 

4 Silverman and Ito (2010) evaluated associations between 2-day rolling average O3

concentrations39 5 and asthma hospital admissions in New York City from 1999 to 2006 (a time 

6 period when the study area would have violated the current standard, Appendix 3-D). As part of 

7 their analysis, the authors evaluated the shape of the concentration-response relationship for O3

8 using a co-pollutant model that included PM2.5 (reprinted in Figure 3-4, below). Based on their 

9 analyses, Silverman and Ito (2010) concluded a linear relationship between O3 and hospital 

10 admissions is a reasonable approximation of the concentration-response function throughout 

11 much of the range of ambient O3 concentrations. Based on visual inspection of Figure 3-4 below

12 (Figure 3 from published study), we note that confidence in the reported concentration-response 

13 relationship is highest for area-wide average O3 concentrations around 40 ppb (i.e., near the 

14 reported median of 41 ppb), and decreases notably for concentrations at and below about 20 ppb. 

15

 

39 2-day rolling averages of 8-hour daily maximum O3 concentrations were calculated throughout the study period, 

averaged across study monitors. 
3-67

1

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

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

and Ito, 2010).40 4

5 In considering the concentration-response function presented by Silverman and Ito (2010) 

6 within the context of the adequacy of the current standard, we evaluate the extent to which the O3

7 concentrations contributing to various portions of the function would likely have been allowed 

8 by the current standard. In doing so, we recognize that true design values cannot be identified for 

9 the subsets of air quality data contributing to various portions of the concentration-response 

function.41 10 Therefore, to use this analysis to inform our consideration of the adequacy of the 

11 current standard we evaluate the extent to which the concentration-response function indicates a 

12 relatively high degree of confidence in the reported health effect association on days when all 

13 monitored 8-hour O3 concentrations were below 75 ppb (Table 3-4, below). This approach can 

14 provide insight into the extent to which the reported O3 health effect association is present when 

15 all monitored O3 concentrations are below the level of the current standard. 

16 Based on the information in Table 3-4 below, when 2-day averaged O3 concentrations 

17 ranged from 36 to 45 ppb (i.e., around the median, where confidence intervals are narrowest), 

18 there was 1 day (out of 236) with at least one monitor recording an 8-hour daily maximum O3

 

40This figure was also reprinted in the ISA (U.S. EPA, 2013; Figure 6-16). 

41As discussed above, O3 design values are calculated using all data available from a monitor. 
3-68

1 concentration above the level of the current standard (approximately 0.4% of days). When 2-day 

2 averaged O3 concentrations ranged from 26 to 45 ppb (i.e., extending to concentrations below the 

3 median, but still above the concentrations where confidence intervals widen notably), there were 

4 4 days (out of 816) with at least one monitor recording an 8-hour daily maximum O3

5 concentration above the level of the current standard (approximately 0.5% of days). Thus, on 

6 over 99% of the days when area-wide “averaged” O3 concentrations were between 26 and 45 

7 ppb, the highest 8-hour daily maximum O3 concentrations were below 75 ppb. For comparison, 

the annual 4th highest 8-hour daily maximum O3 concentration generally corresponds to the 98th 8

or 99th 9 percentile of the seasonal distribution, depending on the length of the O3 season. 

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

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

12 study area evaluated by Silverman and Ito, 2010) 

13

14 In a separate study, Strickland et al. (2010) evaluated associations between 3-day rolling 

average O3 concentrations42 15 and asthma hospital admissions in Atlanta during the warm season 

16 from 1994 to 2004 (a time period when the study area would have violated the current standard, 

17 Appendix 3-D). As part of this analysis, Strickland et al. (2010) evaluated the concentration18 response relationship for O3 and pediatric asthma emergency department visits. The authors 

19 reported the shape of the concentration-response function to be approximately linear with no 

20 evidence of a threshold when 3-day averaged 8-hour daily maximum O3 concentrations were 

21 approximately 30 to 80 ppb (Figure 3-5 below and U.S. EPA, 2013, Figure 6-18). Figure 3-5 

22 below illustrates that the confidence intervals around the concentration-response function are 

 

42 Three-day rolling averages of population-weighted 8-hour daily maximum O3 concentrations were calculated 

throughout the study period (Strickland et al., 2010). 

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)

Min 16 27 30 36 41 45 52 58 62

5th 20 28 34 39 44 49 56 61 67

25th 26 31 37 43 47 53 61 67 72

50th 29 35 42 46 51 59 64 71 76

75th 35 38 46 50 55 64 69 78 85

95th 39 50 54 60 63 74 80 85 96

98th 41 54 60 70 68 82 85 93 99

99th 42 55 67 72 71 87 87 94 103

Max 42 59 80 75 79 100 97 96 108

Days > 75 ppb 0 0 1 0 1 2 9 15 20

2-day moving average across monitors (ppb)

8-hr max from 

highest monitor 

during those 2 days
3-69

1 narrowest around the study mean (i.e., 55 ppb), and that these confidence intervals do not widen 

2 notably for “averaged” O3 concentrations as low as about 30 ppb. 

3

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

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

concentrations (reprinted from Strickland et al., 2010).43 6

7 Similar to the study by Silverman and Ito (2010), we consider the extent to which the 

8 reported concentration-response function indicates a relatively high degree of confidence in 

9 health effect associations on days when all monitored 8-hour O3 concentrations are below 75 ppb 

(Tabe 3-5, below).44 10 In considering the information presented in Table 3-5, we first note that 

11 when 3-day averaged O3 concentrations were in the range of the mean (i.e., 51 to 60 ppb), there 

12 were 77 days (out of 516; 14.9%) with at least one monitor recording an 8-hour daily maximum 

13 O3 concentration above the level of the current standard. In contrast, during the 519 days when 

14 averaged O3 concentrations were in the lower portion of the distribution where study authors 

15 indicate relatively high confidence in the reported concentration-response relationship (i.e., 

16 between 31 and 45 ppb), there were 4 days with at least one monitor in the study area measuring 

17 an 8-hour daily maximum O3 concentration greater than 75 ppb (approximately 0.8% of days). 

 

43 This figure was also reprinted in the ISA (U.S. EPA, 2013; Figure 6-18). 

44 The study by Strickland et al. (2010) used five monitors. For our evaluation of highest 8-hour daily maximum 

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 3-D). 
3-70

1 Thus, on over 99% of the days when “averaged” O3 concentrations were between 31 and 45 ppb, 

2 all monitors measured 8-hour daily maximum O3 concentrations below 75 ppb. 

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

4 monitors over range of 3-day moving averages of population-weighted 

5 concentrations (for study area evaluated by Strickland et al., 2010) 

6

7 In summary, analyses of air quality data from the study locations evaluated by Silverman 

8 and Ito (2010) and Strickland et al. (2010) indicate a relatively high degree of confidence in 

9 reported statistical associations with respiratory health outcomes on days when all monitors 

10 recorded 8-hour average O3 concentrations of 75 ppb or below. Though these analyses do not 

11 identify true design values, the presence of O3-associated respiratory effects on such days 

12 provides insight into the types of health effects that could occur in locations that meet the current 

13 standard. 

14 There are several important uncertainties that are specifically related to our analyses of 

15 distributions of O3 air quality in the study locations evaluated by Silverman and Ito (2010) and 

16 Strickland et al. (2010). Although these studies report health effect associations with two-day 

17 (Silverman and Ito) and three-day (Strickland) averages of daily O3 concentrations, it is possible 

18 that the respiratory morbidity effects reported in these studies were also at least partly 

19 attributable to the days immediately preceding these two- and three-day periods. In support of 

20 this possibility, Strickland et al. reported positive and statistically significant associations with 

21 emergency department visits for multiple lag periods, including lag periods exceeding three days. 

22 Our analysis of highest monitored concentrations focuses on two- and three- day periods, as used 

23 in the published study to generate concentration-response functions. This could have important 

24 implications for our interpretation of the reported concentration-response functions if a 2-day 

25 period with no monitors measuring 8-hour concentrations at or above 75 ppb is immediately 

26 preceded by one or more days with monitors that do exceed 75 ppb. Although we do not know 

27 the extent to which O3 concentrations on a larger number of days could have contributed to 

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)

Min 27 24 33 30 36 38 49 52 57 67 69

5th 29 33 36 45 48 54 60 64 69 75 85

25th 36 40 45 51 56 63 68 73 79 85 95

50th 38 44 50 58 62 68 73 78 86 92 101

75th 44 49 54 62 70 74 81 85 96 101 108

95th 53 56 70 74 81 84 95 95 104 109 120

98th 59 59 75 84 86 93 100 98 107 110 124

99th 66 61 78 89 88 95 103 102 108 112 124

Max 70 64 83 99 97 107 106 122 129 120 138

Days > 75 0 0 2 2 10 24 53 80 89 87 87

8-hr max 

from 

highest 

monitor 

during 

3-day moving average across monitors (ppb)
3-71

1 reported health effect associations, we note this as a potentially important uncertainty in our 

2 consideration of concentration-response functions within the context of the current standard. 

3 In addition, an important uncertainty that applies to epidemiologic studies in general is 

4 the extent to which reported health effects are caused by exposures to O3 itself, as opposed to 

5 other factors such as co-occurring pollutants or other pollutant mixtures. Although both of the 

6 studies evaluated above reported health effect associations in co-pollutant models, this 

7 uncertainty becomes an increasingly important consideration as health effect associations are 

8 evaluated at lower ambient O3 concentrations (i.e., resulting in lower exposure concentrations). 

9 One approach to considering the potential importance of this uncertainty in 

10 epidemiologic studies is to evaluate the extent to which there is coherence with the results of 

11 experimental studies (i.e., in which the study design dictates that exposures to O3 itself are 

12 responsible for reported effects). Therefore, in further considering uncertainties associated with 

13 the above air quality analyses for the study areas evaluated by Silverman and Ito (2010) and 

14 Strickland et al. (2010), we evaluate the following question: 

15 To what extent is there coherence between evidence from controlled human 

16 exposure studies and epidemiologic studies supporting the occurrence of O3-

17 attributable respiratory effects when 8-hour daily maximum ambient O3 

18 concentrations are at or below 75 ppb? 

19 As summarized above and as discussed in the ISA (U.S. EPA, 2013, section 6.2), 

20 controlled human exposure studies demonstrate the occurrence of respiratory effects in an 

21 appreciable percentage of healthy adults following single short-term exposures to O3

22 concentrations as low as 60 ppb. In addition, as O3 exposure concentrations exceed 60 ppb: 1) 

23 effects in healthy adults become larger and more serious; 2) a broader range of effects are 

24 observed in a greater percentage of exposed individuals; and 3) effects are reported more 

25 consistently across studies. Thus, as the potential increases for exposures to O3 concentrations 

26 approaching or exceeding 60 ppb, our confidence increases that reported respiratory health 

27 effects could be caused by exposures to the ambient O3 concentrations present in study locations. 

28 In considering coherence with results of controlled human exposure studies in this way, it 

29 is important to note the relative lack of experimental data for exposure concentrations below 60 

30 ppb. It is possible that lower exposure concentrations can result in respiratory effects serious 

31 enough to lead to hospital admissions or emergency department visits, particularly in at-risk 

32 populations such as children and asthmatics. Thus, although we consider coherence between 

33 epidemiologic and controlled human exposure study results, we also acknowledge that an O3

34 exposure concentration of 60 ppb does not constitute a bright line below which we are confident 

35 that effects no longer occur, particularly for at-risk groups. 
3-72

1 As discussed above, for the study by Silverman and Ito (2010), 26 to 45 ppb represents 

2 the lower end of the range of “averaged” daily maximum 8-hour concentrations over which the 

3 study indicates a relatively high degree of confidence in the statistical association with 

4 respiratory hospital admissions (and for which virtually all monitored concentrations were 75 

5 ppb or below). As averaged concentrations increase from 26 to 45 ppb, the number of days with 

6 maximum monitored concentrations exceeding 60 ppb increases dramatically (Table 3-4, 

above).45 7 For example, of the 178 days with area-wide average daily maximum 8-hour 

8 concentrations from 26 to 30 ppb, only about 2% had any monitors recording ambient 

9 concentrations of 60 ppb or greater. In contrast, of the 196 days with area-wide average 

10 concentrations from 41 to 45 ppb, about half had at least one monitor recording an ambient 

11 concentration near or above 60 ppb, with monitors on some days measuring concentrations 

12 greater than 80 ppb. Thus, as averaged concentrations approach 45 ppb there is an increasing 

13 likelihood that at least some portion of the study population could have been exposed to O3

14 concentrations approaching or exceeding those reported in controlled human exposure studies to 

15 cause respiratory effects in healthy adults. 

16 For the study by Strickland et al (2010), “averaged” concentrations from 30 to 45 ppb

17 represent the lower end of the range of concentrations over which the study indicates a relatively 

18 high degree of confidence in the statistical association with respiratory emergency department 

19 visits (and for which virtually all monitored concentrations were 75 ppb or below). Similar to the 

20 study area evaluated by Silverman and Ito, as 8-hour area-wide average O3 concentrations 

21 approach 45 ppb, maximum monitored concentrations exceed 60 ppb more frequently. On most 

22 days contributing to averaged O3 concentrations from 41 to 45 ppb, maximum monitor 

23 concentrations were near or above 60 ppb. On a small number of these days, maximum 

24 monitored concentrations were greater than 80 ppb. Therefore, similar to the study by Silverman 

25 and Ito (2010), at least some portion of the study population on these days are likely to have been 

26 exposed to O3 concentrations exceeding those reported in controlled human exposure studies to 

27 cause respiratory effects in some healthy adults. 

28 Thus, consideration of distributions of individual monitored concentrations, in 

29 conjunction with the results of O3 controlled human exposure studies, supports the role of 

30 ambient O3 concentrations in eliciting the reported respiratory hospital admissions and 

31 emergency department visits. Specifically, these analyses support the occurrence of O3-

32 attributable hospital admissions and emergency department visits on days when virtually all 

33 monitors measure 8-hour ambient concentrations below 75 ppb.

 

45Though, 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. 
3-73

1 In further evaluating O3 concentration-response relationships within the context of the 

2 adequacy of the current standard, we note that some epidemiologic studies report health effect 

3 associations for air quality distributions restricted to ambient pollutant concentrations below one 

4 or more predetermined cut points. Such “cut point” analyses can provide information on the 

5 magnitude and statistical precision of effect estimates for defined distributions of ambient 

6 concentrations, which may in some cases include distributions that would meet the current 

7 standard. Therefore, we next consider the following question: 

8 To what extent do cut-point analyses from epidemiologic studies report health effect 

9 associations at ambient O3 concentrations lower than previously identified or that 

10 would likely meet the current standard?

11 By considering the magnitude and statistical significance of effect estimates for restricted 

12 air quality distributions, cut-point analyses can provide insight into the extent to which health 

13 effect associations are driven by ambient concentrations above the cut point, versus 

14 concentrations below the cut point. For studies that evaluate multiple cut points, these analyses 

15 can provide insights into the magnitude and statistical precision of health effect associations for 

16 different portions of the distribution of ambient concentrations, including insights into the 

17 ambient concentrations below which uncertainty in reported associations becomes notably 

18 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.46 19

20 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 concentrations47 21 and total (non22 accidental) mortality in 98 U.S. cities from 1987 to 2000. Based on the full distributions of 

23 ambient O3 concentrations in study cities, the large majority of the NMMAPS cities would have 

24 violated the current standard during the study period (Appendix 3-D). However, Bell et al. 

25 (2006) also reported health effect associations in a series of cut-point analyses, with effect 

26 estimates based only on the subsets of days contributing to “averaged” O3 concentrations below 

27 cut points ranging from 5 to 60 ppb (see Figure 2 in Bell et al., 2006). The lowest cut-point for 

28 which the association between O3 and mortality was reported to be statistically significant was 

29 30 ppb (based on visual inspection of Figure 2 in published study). As with the studies by 

30 Silverman and Ito (2010) and Strickland et al. (2010), we consider what these cut point analyses 

 

46As 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. 

47Two-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. 
3-74

1 indicate with regard to the potential for health effect associations to extend to ambient O3

2 concentrations likely to be allowed by the current O3 NAAQS. 

3 We attempted to recreate the subsets of air quality data used in the cut point analyses 

4 presented by Bell et al. (2006). In doing so, we applied the criteria described in the published 

5 study to generate air quality subsets corresponding to those defined by the cut points evaluated 

by study authors.48 6 From the days with averaged O3 concentrations below each cut point, we 

identified 3-year averages of annual 4th 7 highest 8-hour daily maximum O3 concentrations in each 

study area. We then compared these 4

th 8 highest O3 concentrations to the level of the current 

9 standard in order to provide insight into the extent to which the air quality distributions included 

10 in various cut point analyses would likely have met the current standard. 

11 We particularly focus on the lowest cut-point for which the association between O3 and 

12 mortality was reported to be statistically significant (i.e., 30 ppb, as noted above). Based on the 

13 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 14 highest 8-hour daily maximum O3

15 concentration at or below 75 ppb over the entire study period. For the 35 ppb cut point, which 

16 also resulted in a statistically significant association with mortality, 68% of study areas had 3-

year averages of annual 4th 17 highest 8-hour daily maximum O3 concentration at or below 75 ppb. 

18 This suggests that the large majority of air quality distributions that provided the basis for 

19 positive and statistically significant associations with mortality (i.e., for the 30 and 35 ppb cut 

20 points) would likely have met the current O3 standard. For higher cut points, all of which also 

21 resulted in statistically significant associations with mortality, the majority of study cities had 3-

year averages of annual 4th 22 highest 8-hour daily maximum concentrations greater than 75 ppb. 

23

 

48We 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. 
3-75

Table 3-6. Number of study cities with 4th 1 highest 8-hour daily maximum concentrations 

2 greater than 75 ppb, for various cut-point analyses presented in Bell et al. 

3 (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%)

4

5 In addition to the uncertainties noted above for our analysis of the single-city studies by 

6 Silverman and Ito (2010) and Strickland et al. (2010) (e.g., attributing effects specifically to air 

7 quality included in various subsets), an important uncertainty related to this analysis is that we 

8 were unable to obtain the air quality data used to generate the cut-point analyses in the study 

9 published by Bell et al. (2006). Therefore, as noted above, we generated 2-day averages of 24-

10 hour O3 concentrations in study locations using the air quality data available in AQS, combined 

11 with the published description of study area definitions. In doing so, we did not recreate the 

12 trimmed means used by Bell. An important uncertainty in this approach is the extent to which we 

13 were able to appropriately recreate the cut-point analyses in the published study. 

14 The ISA also notes important uncertainties inherent in multicity studies that evaluate the 

15 potential for thresholds to exist, as was done in the study by Bell et al. (2006). Specifically, the 

16 ISA highlights the regional heterogeneity in O3 health effect associations as a factor that could 

17 obscure the presence of thresholds, should they exist, in multicity studies (U.S. EPA, 2013, 

18 sections 2.5.4.4 and 2.5.4.5). The ISA notes that community characteristics (e.g., activity 

19 patterns, housing type, age distribution, prevalence of air conditioning) could be important 

20 contributors to reported regional heterogeneity (U.S. EPA, 2013, section 2.5.4.5). Given this 

21 heterogeneity, the ISA concludes that “a national or combined analysis may not be appropriate to 

22 identify whether a threshold exists in the O3-mortality C-R relationship” (U.S. EPA, 2013, p. 2-

23 33). This represents an important source of uncertainty when characterizing our confidence in 

24 reported concentration-response relationships over distributions of ambient O3 concentrations, 

25 based on multicity studies. This uncertainty becomes increasingly important when interpreting 

26 concentration-response relationships at lower ambient O3 concentrations, particularly those 

27 concentrations corresponding to portions of distributions where data density decreases notably. 

28 3.1.4.3 Concentrations in Epidemiologic Studies – “Long-term” Metrics

29 We next consider the extent to which epidemiologic studies employing longer-term 

30 ambient O3 concentration metrics inform our understanding of the air quality conditions 
3-76

1 associated with O3-attributable health effects, and specifically inform consideration of the extent 

2 to which such effects could occur under air quality conditions meeting the current standard. 

3 Unlike for the studies of short-term O3 discussed above, the available U.S. and Canadian 

4 epidemiologic studies evaluating long-term ambient O3 concentration metrics have not been 

5 conducted in locations likely to have met the current 8-hour O3 standard during the study period 

6 (Appendix 3-D). Therefore, although these studies contribute to our understanding of health 

7 effects associated with long-term or repeated exposures to ambient O3 (as summarized in section 

8 3.1.2 above), consideration of study area design values does not inform our consideration of the 

9 extent to which those health effects may be occurring in locations that met the current standard. 

10 In further considering epidemiologic studies of long-term O3 concentrations, we also 

11 evaluate the extent to which concentration-response functions have been reported over 

12 distributions of ambient concentrations, and what those functions can tell us about health effect 

13 associations with O3 concentrations likely to be allowed by the current standard. Specifically, we 

14 consider the following question: 

15 To what extent do confidence intervals around concentration-response functions 

16 indicate O3-associated health outcomes at ambient concentrations meeting the 

17 current O3 standard? 

18 The ISA identifies a single epidemiologic study reporting confidence intervals around a 

19 concentration-response function for “long-term” O3 concentrations and respiratory mortality 

20 (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 concentrations49 21 ranged from 33 to 

22 104 ppb, there was no statistical deviation from a linear concentration-response relationship 

23 between O3 and respiratory mortality across 96 U.S. cities (U.S. EPA, 2013, section 7.7). 

24 However, the authors reported “limited evidence” for an effect threshold at an O3 concentration 

25 of 56 ppb (p=0.06). Visual inspection of this concentration-response function (Figure 3-6) 

26 confirms the possibility of an inflection point just below 60 ppb, which is close to the median 

27 concentration across cities (i.e., 57 ppb). 

 

49Jerrett 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. 
3-77

1

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

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

4 We consider the extent to which this concentration-response function indicates 

5 confidence in the reported health effect association at various ambient O3 concentrations. In 

6 doing so, we note the following: (1) most of the study cities had O3 concentrations above 53.1

7 ppb (i.e., the upper bound of the first quartile), accounting for approximately 72% of the 

8 respiratory deaths in the cohort (Table 2 in Jerrett et al. 2009); (2) confidence intervals widen 

9 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.50 10 In considering this 

11 information, we conclude that the analysis reported by Jerrett indicates the greatest confidence in 

12 the linear concentration-response function for “long-term” O3 concentrations in the upper three 

13 quartiles (i.e., above about 53 ppb). 

14 Based on information in the published study (Figure 1 in Jerrett et al., 2009), we 

15 identified 72 of the 96 study cities as having ambient O3 concentrations in the highest three 

16 quartiles (Appendix 3-D). As noted above, these 72 cities account for approximately 72% of the 

17 respiratory deaths in the cohort (Table 2 in Jerrett et al. 2009). Of these 72 cities, 71 had 3-year 

averages of annual 4th 18 highest 8-hour daily maximum O3 concentrations above 75 ppb (Appendix 

19 3-D). Thus, the current 8-hour NAAQS would have been violated during the study period in 

20 virtually all of the study cities that contribute to the range of long-term O3 concentrations over 

21 which we have the greatest confidence in the reported relationship with respiratory mortality.

22 Thus, while the study by Jerrett et al. (2009) contributes to our understanding of health effects 

23 associated with ambient O3 (as summarized in section 3.1.2 above), it is less informative 

 

50The ISA does not reach conclusions regarding the potential for a threshold in the association between “long-term” 

O3 concentrations and respiratory mortality. 
3-78

1 regarding the extent to which those health effects may be occurring under air quality conditions 

2 allowed by the current standard. 

3 3.1.5 Public Health Implications 

4 In this section, we address the public health implications of O3 exposures with respect to 

5 the factors that put populations at increased risk from exposures (section 3.1.5.1), the size of at6 risk populations (section 3.1.5.2), and the potential effects of averting behavior on reducing O3

7 exposures and associated health effects (section 3.1.5.3). Providing appropriate public health 

8 protection requires consideration of the factors that put populations at greater risk from O3

9 exposure. In order to estimate potential overall for public health impacts, it is important to 

10 consider not only the adversity of the health effects, but also the populations at greater risk and 

11 potential behaviors that may reduce exposure. 

12 3.1.5.1 At-Risk Populations 

13 In this section we address the following question: 

14 To what extent does the currently available scientific evidence expand our 

15 understanding of at-risk populations? 

16 The currently available evidence expands our understanding of populations that were 

17 identified to be at greater risk of O3-related health effects at the time of the last review (i.e., 

18 people who are active outdoors, people with lung disease, children and older adults and people 

19 with increased responsiveness to O3) and supports the identification of additional factors that 

20 may lead to increased risk (U.S. EPA 2006, section 3.6.2; U.S. EPA, 2013, chapter 8). 

21 Populations and lifestages may be at greater risk for O3-related health effects due to factors 

22 contribute to their susceptibility and/or vulnerability to ozone. The definitions of susceptibility 

23 and vulnerability have been found to vary across studies, but in most instances “susceptibility” 

24 refers to biological or intrinsic factors (e.g., lifestage, sex, preexisting disease/conditions) while 

25 “vulnerability” refers to non-biological or extrinsic factors (e.g., socioeconomic status [SES]) 

26 (U.S. EPA, 2013, p. 8-1; U.S. EPA, 2010c, 2009d). In some cases, the terms “at-risk” and 

27 “sensitive” have been used to encompass these concepts more generally. In the ISA and this PA, 

28 “at-risk” is the all-encompassing term used for groups with specific factors that increase their

29 risk of O3-related health effects. Further discussion of at-risk populations can be found in 

30 Appendix 3C below. 

31 There are multiple avenues by which groups may experience increased risk for O3-

induced health effects. A population or lifestage51 32 may exhibit greater effects than other 

 

51 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

1 populations or lifestages exposed to the same concentration or dose, or they may be at greater 

2 risk due to increased exposure to an air pollutant (e.g., time spent outdoors). A group with 

3 intrinsically increased risk would have some factor(s) that increases risk through a biological 

4 mechanism and, in general, would have a steeper concentration-risk relationship, compared to 

5 those not in the group. Factors that are often considered intrinsic include asthma, genetic 

6 background, and lifestage. A group of people could also have extrinsically increased risk, which 

7 would be through an external, non-biological factor, including, for example, socioeconomic 

8 status (SES) and diet. Some groups are at risk of increased internal dose at a given exposure 

9 concentration, for example, because of breathing patterns. This category would include people 

10 who work or exercise outdoors. Finally, there are those who might be placed at increased risk for 

11 experiencing greater exposures by being exposed at higher concentrations. This would include, 

12 for example, groups of people with greater exposure to ambient O3 due to less availability or use 

13 of home air conditioners such that they are more likely to be in locations with open windows on 

14 high ozone days. Some groups may be at increased risk of O3-related health effects through a 

15 combination of factors. For example, children tend to spend more time outdoors when O3 levels 

16 are high, and at higher levels of activity than adults, which leads to increased exposure and dose, 

17 and they also have biological, or intrinsic, risk factors (e.g., their lungs are still developing) 

18 (U.S. EPA, 2013, Chapter 8). An at-risk population or lifestage is more likely to experience 

19 adverse health effects related to O3 exposures and/or, develop more severe effects from exposure 

20 than the general population.

21 Based on the currently available evidence, the at-risk populations for O3- related health 

22 effects are based on factors that include: asthma, lifestages (children and older adults), genetic 

23 variability, dietary factors, and working outdoors (U.S. EPA, 2013, section 8.5, Table 8-6). This 

24 conclusion is supported by consistency in findings across studies and evidence of coherence in 

25 results from different scientific disciplines. The current evidence is suggestive of a potential for 

26 three other factors to influence risk of O3-related health effects. The evidence suggests that 

27 women may be at greater risk than men, groups with low SES or living in neighborhoods with 

28 low SES may be at greater risk than other socioeconomic groups, and obesity may be a potential 

29 risk factor. Further studies are needed, however, on these factors. Overall, the factors most 

30 strongly supported as contributing to increased risk of O3-related effects are related to asthma, 

31 lifestage (children and older adults), genetic variability, dietary factors, and working outdoors 

32 (U.S. EPA, 2013; chapter 8).

33 In summary, the evidence available in this review supports the identification of the 

34 following populations and lifestages as having increased risk for O3-related health effects, based 

35 on consistency in findings across studies and evidence of coherence in results from different 
3-80

1 scientific disciplines (U.S. EPA, 2013, section 8.5): individuals with certain genotypes, 

2 individuals with asthma, younger and older age groups, individuals with reduced intake of 

3 certain nutrients, and outdoor workers. Multiple genetic variants have been observed in 

4 epidemiologic and controlled human exposure studies to affect the risk of O3-related respiratory 

5 outcomes and support is provided by toxicological studies of genetic factors (U.S. EPA, 2013, 

6 section 8.1). Asthma has been recognized in past reviews and continues to be well established in 

7 this review as a risk factor for O3-related health effects based on multiple lines of evidence, 

8 included controlled human exposure and toxicological studies in animal models, as well as some 

9 evidence from epidemiologic studies. This evidence supports our understanding of the 

10 biological (intrinsic) factors that put individuals with asthma at greater risk than other groups 

11 (U.S. EPA, 2013, section 8.2.2). As noted above, some extrinsic (exposure-related) and intrinsic 

12 factors contribute to the identification of children as an at-risk lifestage. Children have higher 

13 exposure and dose due to increased time spent outdoors and ventilation rate, their lungs are still 

14 developing, and they are more likely than adults to have asthma (U.S. EPA, 2013, section 

15 8.3.1.1). Older adults may also withstand greater O3 exposure and not seek relief as quickly as 

16 younger adults. Multiple epidemiologic, controlled human exposure and toxicological studies 

17 reported that diets deficient in vitamins E and C are associated with risk of O3-related health 

18 effects for all lifestages. Previous studies have shown that increased exposure to O3 due to 

19 outdoor work leads to increased risk of O3-related health effects and it is clear that outdoor 

20 workers have higher exposures, and possibly greater internal doses, of O3, which may lead to 

21 increased risk of O3-related health effects (U.S. EPA, 2013, section 8.5). 

22 In some cases, it is difficult to determine a factor that results in increased risk of effects. 

23 For example, previous assessments have included controlled human exposure studies in which 

24 some healthy individuals demonstrate greater O3-related health effects compared to other healthy 

25 individuals. Intersubject variability has been observed for lung function decrements, 

26 symptomatic responses, pulmonary inflammation, AHR, and altered epithelial permeability in 

27 healthy adults exposed to O3 and these results tend to be reproducible within a given individual 

28 over a period of several months indicating differences in the intrinsic responsiveness. In many 

29 cases the reasons for the variability is not clear. This may be because one or some of the factors 

30 described above have not been evaluated in studies, or it may be that additional, unidentified 

31 factors influence individual responses to O3 (U.S. EPA, 2013, section 8.5). 

32 As discussed in chapter 8 of the ISA, and further in Appendix 3C below, the challenges 

33 and limitations in evaluating the factors that can increase risk for experiencing O3-related health 

34 effects may contribute to a lack of information about the factors that may increase risk from O3

35 exposures. This lack of information may contribute to conclusions that evidence for some 
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1 factors, such as sex, SES, and obesity provided “suggestive” evidence of increased risk, or that 

2 for a number of factors the evidence was inadequate to draw conclusions about potential increase 

3 in risk of effects. Overall, the factors most strongly supported as contributing to increased risk of 

4 populations for experience O3-related effects were related to asthma, lifestage (children and older 

5 adults), genetic variability, dietary factors, and working outdoors. 

6 3.1.5.2 Size of At-Risk Populations and Lifestages in the United States 

7 One consideration in the assessment of potential public health impacts is the size of 

8 various population groups for which there is adequate evidence of increased risk for health 

9 effects associated with O3-related air pollution exposure. The factors for which the ISA judged 

10 the evidence to be “adequate” with respect to contributing to increased risk of O3-related effects 

11 among various populations and lifestages included: asthma; childhood and older adulthood; diets 

12 lower in vitamins C and E; certain genetic variants and, working outdoors (EPA, 2013, section 

13 8.5). 

14 With regard to asthma, Table 3-12 below summarizes information on the prevalence of 

15 current asthma by age in the U.S. adult population in 2010 (Schiller et al. 2012; children - Bloom 

16 et al., 2011). Individuals with current asthma constitute a fairly large proportion of the 

17 population, including more than 25 million people. Asthma prevalence tends to be higher in 

18 children than adults. 

19 Within the U.S., approximately 8.2% of adults have reported currently having asthma 

20 (Schiller et al., 2012) and 9.5% of children have reported currently having asthma (Bloom et al., 

21 2011). Table 3-12 below provides more detailed information on prevalence of asthma by age in 

22 the U.S. 

23
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1 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)

2

3 With regard to lifestages, based on U.S. census data from 2010 (Howden and Meyer, 

4 2011), about 74 million people, or 24% of the U.S. population, are under 18 years of age and 

5 more than 40 million people, or about 13% of the U.S. population, are 65 years of age or older. 

6 Hence, a large proportion of the U.S. population, more than 33%, is included in age groups that 

7 are considered likely to be at increased risk for health effects from ambient O3 exposure. 

8 With regard to dietary factors, no statistics are available to estimate the size of an at-risk

9 population based on nutritional status. 

10 With regard to outdoor workers, in 2010 approximately 11.7% of the total number of 

11 people (143 million people) employed, or about 16.8 million people, worked outdoors one or 

more day per week (based on worker surveys).52 12 Of these approximately 7.4% of the workforce, 

13 or about 7.8 million people, worked outdoors three or more days per week. 

14 The health statistics data illustrate what is known as the “pyramid” of effects. At the top 

15 of the pyramid, there are approximately 2.5 million deaths from all causes per year in the U.S. 

16 population, with about 250 thousand respiratory-related deaths (CDC-WONDER, 2008). For 

17 respiratory health diseases, there are nearly 3.3 million hospital discharges per year (HCUP, 

18 2007), 8.7 million respiratory ED visits (HCUP, 2007), 112 million ambulatory care visits 

 

52 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/ 
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1 (Woodwell and Cherry, 2004), and an estimated 700 million restricted activity days per year due 

2 to respiratory conditions (Adams et al., 1999). Combining small risk estimates with relatively 

3 large baseline levels of health outcomes can result in quite large public health impacts. Thus, 

4 even a small percentage reduction in O3 health impacts on cardiopulmonary diseases would 

5 reflect a large number of avoided cases. 

6 3.1.5.3 Averting Behavior 

7 The activity pattern of individuals is an important determinant of their exposure (ISA, 

8 U.S. EPA, 2013, section 4.4.1). Variation in O3 concentrations among various 

9 microenvironments means that the amount of time spent in each location, as well as the level of 

10 activity, will influence an individual’s exposure to ambient O3. Activity patterns vary both 

11 among and within individuals, resulting in corresponding variations in exposure across a 

12 population and over time. Individuals can reduce their exposure to O3 by altering their behaviors, 

13 such as by staying indoors, being active outdoors when air quality is better, and by reducing their 

14 activity levels or reducing the time being active outdoors on high-O3 days (U.S. EPA, 2013, 

15 section 4.4.2). The evidence in this topic area, while not addressed in the 2006 AQCD, is 

16 evaluated in the ISA for this review. 

17 The widely reported Air Quality Index (AQI) conveys advice to the public, and 

18 particularly at-risk populations, on reducing exposure on days when ambient levels of common 

19 air pollutants are elevated (www.airnow.gov). The AQI describes the potential for health effects 

20 from O3 (and other individual pollutants) in six color-coded categories of air-quality, ranging 

21 from Good (green), Moderate (yellow), Unhealthy for Sensitive Groups (orange), Unhealthy 

22 (red), and Very Unhealthy (purple), and Hazardous (maroon). Levels in the unhealthy ranges 

23 (i.e., Unhealthy for Sensitive Groups and above) come with recommendations about reducing 

24 exposure. Forecasted and actual AQI values for O3 are reported to the public during the O3

25 season. The AQI advisories explicitly state that children, older adults, people with lung disease, 

26 and people who are active outdoors, may be at greater risk from exposure to O3. People are 

27 advised to reduce exposure depending on the predicted O3 levels and the likelihood of risk. This 

28 advice includes being active outdoors when air quality is better, and reducing activity levels or 

29 reducing the time being active outdoors on high-O3 days. Staying indoors to reduce exposure is 

30 not recommended until air quality reaches the Very Unhealthy or Hazardous categories. 

31 Evidence of individual averting behaviors in response to AQI advisories has been found 

32 in several studies, including activity pattern and epidemiologic studies, especially for the at-risk 

33 populations, such as children, older adults, and people with asthma, who are targeted by the 

34 advisories. Such effects are less pronounced in the general population, possibly due to the 

35 opportunity cost of behavior modification. Epidemiologic evidence from a study (Neidell and 
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1 Kinney, 2010) conducted in the 1990’s in Los Angeles, CA reports increased asthma hospital 

2 admissions among children and older adults when O3 alert days (1-hour max O3 concentration 

3 >200 ppb) were excluded from the analysis of daily hospital admissions and O3 concentrations 

4 (presumably thereby eliminating averting behavior based on high O3 forecasts). The lower rate 

5 of admissions observed when alert days were included in the analysis suggests that estimates of 

6 health effects based on concentration-response functions that do not account for averting 

7 behavior may be biased towards the null (U.S. EPA, 2013, section 4.4.2). 

8 3.2 AIR QUALITY-, EXPOSURE-, AND RISK-BASED CONSIDERATIONS 

9 In order to inform judgments about the public health impacts of O3-related health 

10 effects, the second draft HREA has developed and applied models to estimate human exposures 

11 to O3 and O3-associated health risks across the United States, with a specific focus on urban case 

study areas (U.S. EPA, 2014).53 12 The second draft HREA uses photochemical modeling to adjust 

13 air quality from the 2006-2010 O3 seasons to just meet the current and alternative standards for 

the 2006-2008 and 2008-2010 periods.54 14 In this section, staff considers estimates of short-term 

15 O3 exposures and estimates of health risks associated with short- and long-term O3 exposures, for 

16 air quality adjusted to just meet the current O3 standard. In section 3.2.1, we consider the 

17 implications for exposure and risk estimates of the approach used in the second draft HREA to 

18 adjust air quality. Sections 3.2.2 and 3.2.3 discuss our exposure-based and risk-based 

19 considerations, respectively. In these sections we specifically consider the following question: 

20 What are the nature and magnitude of O3 exposures and health risks remaining 

21 upon adjusting recent air quality to just meet the current O3 standard, and what are 

22 the important uncertainties associated with those exposure and risk estimates?

23 3.2.1 Consideration of the Adjusted Air Quality Used in Exposure and Risk 

24 Assessments 

25 In the first draft HREA for this review, as in the last review, the EPA relied upon 

26 quadratic rollback to adjust hourly O3 concentrations in urban case study areas to just meet the 

27 current O3 standard (U.S. EPA, 2012b). Although the quadratic rollback method reproduces 

28 historical patterns of air quality changes better than some alternative methods, it relies on 

 

53 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 second draft 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. 

54 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. 
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1 statistical relationships without explicitly accounting for atmospheric chemistry and precursor 

2 emissions (U.S. EPA, 2014, Chapter 4). An important drawback of the quadratic rollback 

3 approach, recognized in the first draft HREA (U.S. EPA, 2012b), is that it forces all monitors in 

4 an assessment area to exhibit the same response when air quality is adjusted. It does not allow for 

5 the spatial or temporal heterogeneity in responses that result from the non-linear atmospheric 

6 chemistry that influences ambient O3 concentrations (U.S. EPA, 2014, Chapter 4). Because 

7 quadratic rollback does not account for physical and chemical atmospheric processes, or the 

8 sources of emissions precursors that lead to O3 formation, a backstop or “floor” must be used 

9 when applying quadratic rollback to just meet current or alternative standards to ensure that 

10 estimated O3 is not reduced in a manner inconsistent with O3 chemistry, such as to reduce 

11 concentrations below that associated with background sources (U.S. EPA, 2014, Chapter 4). 

12 Consistent with recommendations from the National Research Council of the National 

13 Academies (NRC, 2008), the second draft HREA uses a photochemical model to estimate 

14 sensitivities of O3 to changes in precursor emissions, in order to estimate ambient O3

15 concentrations that would just meet the current and alternative standards (U.S. EPA, 2014, 

Chapter 4).55 16 For the urban case study areas evaluated in the second draft HREA, this model17 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)56 18 were reduced 

to estimate air quality that just meets the current and alternative O3 standards.57 19

20 As discussed in Chapter 4 of the second draft HREA (U.S. EPA, 2014), this approach 

21 models the physical and chemical atmospheric processes that influence ambient O3

22 concentrations. Compared to the quadratic rollback approach, it provides more realistic estimates 

23 of the spatial and temporal responses of O3 to reductions in precursor emissions. These improved 

24 estimates avoid many of the limitations inherent in the quadratic rollback method, including the 

25 requirement that all monitors in an assessment area exhibit the same response upon model 

 

55The second draft 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). 

56Exposure and risk analyses for most urban case study areas focus on reducing NOX emissions alone (NOX

emissions were reduced by about 40 to 85% for the current standard, and up to 95% for alternatives). In most of the 

urban case study areas, the addition of modeled reductions in VOC emissions did not alter the reductions in NOX

emissions required to simulate the current or alternative standards. However, in Chicago and Denver, the addition of 

reductions in VOC emissions allowed for smaller NOX emissions reductions to simulate the current and alternative 

standards. Therefore, exposure and risk analyses for Chicago and Denver focus on reductions in emissions of both 

NOX and VOC (U.S. EPA, 2014, section 4.3.3.1). 

57Although this chapter focuses on the current standard, our overarching considerations regarding model-adjusted 

air quality also apply to alternative standards simulated in the second draft HREA. Alternative standards are 

discussed in chapter 4 of this second draft PA. 
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1 adjustment to the current and/or alternative standards. Because model-adjusted air quality 

2 scenarios are based on reducing only U.S. anthropogenic emissions, this approach also does not 

3 require the specification of background concentrations as a rollback “floor” (U.S. EPA, 2014, 

4 section 4.3.3). 

5 The use of this model-based air quality adjustment approach in the second draft HREA 

6 has important implications for the patterns of ambient O3 concentrations estimated in urban case 

7 study areas. Specifically, in locations and time periods when NOX is predominantly contributing 

8 to O3 formation (e.g., downwind of important NOX sources, where the highest O3 concentrations 

9 often occur), model-based adjustment to the current and alternative standards decreases 

10 estimated ambient O3 concentrations compared to recent monitored concentrations (U.S. EPA, 

11 2014, section 4.3.3.2). In contrast, in locations and time periods when NOX is predominantly 

12 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 low58 13 ), model-based 

14 adjustment increases ambient O3 concentrations compared to recent measured levels (U.S. EPA, 

15 2014, section 4.3.3.2) (Chapter 2, above). 

16 Within urban case study areas, the overall impacts of model-based air quality adjustment 

17 are to reduce relatively high ambient O3 concentrations (i.e., concentrations at the upper ends of 

18 ambient distributions) and to increase relatively low O3 concentrations (i.e., concentrations at the 

19 lower ends of ambient distributions) (U.S. EPA, 2014, section 4.3.3.2, Figures 4-8 to 4-11). 

20 Seasonal means of daily concentrations generally exhibit only modest changes upon model 

21 adjustment, reflecting the seasonal balance between daily decreases and increases in ambient 

22 concentrations (U.S. EPA, 2014, Figures 4-10 and 4-11). The resulting compression in 

23 distributions of ambient O3 concentrations is evident in all of the urban case study areas 

24 evaluated, though the degree of compression varies considerably across areas (U.S. EPA, 2014, 

25 Figures 4-10 and 4-11). 

26 Adjusted patterns of O3 air quality have important implications for exposure and risk 

27 estimates in urban case study areas. Estimates influenced largely by the upper ends of the 

28 distribution of ambient concentrations (i.e., exposures of concern and lung function risk 

29 estimates, as discussed in sections 3.2.2 and 3.2.3.1 below) will decrease with model-adjustment 

30 to the current and alternative standards. In contrast, seasonal risk estimates influenced by the full 

31 distribution of ambient O3 concentrations (i.e., epidemiology-based risk estimates, as discussed 

32 in section 3.2.3.2 below) will either increase or decrease in response to model adjustment, 

 

58Titration is also prominent during time periods when photochemistry is limited, such as at night and on cool, 

cloudy days. 
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1 depending on the balance between the daily decreases in high O3 concentrations and increases in 

low O3 concentrations.59 2

3 We further consider the implications of the spatial and temporal patterns of adjusted air 

4 quality within the context of exposure (section 3.2.2) and risk (section 3.2.3) estimates for O3

5 concentrations adjusted to just meet the current standard. As discussed below (section 3.2.3.2), 

6 these altered patterns are particularly important to consider when interpreting epidemiology7 based risk estimates. 

8 3.2.2 Exposure-Based Considerations 

9 The exposure assessment presented in the second draft HREA (U.S. EPA, 2014) provides 

10 estimates of the number of people exposed to various concentrations of ambient O3, while at 

11 specified exertion levels. The second draft HREA estimates exposures in 15 urban case study 

12 areas for school-age children (ages 5 to 18), asthmatic school-age children, asthmatic adults, and 

13 older adults, reflecting the strong evidence indicating that these populations are potentially at

14 increased risk for O3-attributable effects (EPA, 2013, Chapter 8; section 3.1.2, above). An 

15 important purpose of these exposure estimates is to provide perspective on the extent to which 

16 air quality adjusted to just meet the current O3 NAAQS could be associated with exposures to O3

concentrations reported to result in respiratory effects.60 17 Estimates of such “exposures of 

18 concern” provide perspective on the potential public health impacts of O3-related effects, 

19 including for effects that cannot currently be evaluated in a quantitative risk assessment (e.g., 

20 airway inflammation). 

21 In the absence of large scale exposure studies that encompass the general population, as 

22 well as at-risk populations, modeling is the preferred approach to estimating exposures to O3. 

23 Additionally, the use of exposure modeling facilitates the estimation of exposures resulting from 

24 ambient air concentrations differing from those in exposure studies (e.g., concentrations just 

25 meeting the current standard). In the second draft HREA, population exposures to ambient O3

26 concentrations are estimated using the current version of the Air Pollutants Exposure (APEX) 

27 model. The APEX model simulates the movement of individuals through time and space and 

28 estimates their exposures to a given pollutant in indoor, outdoor, and in-vehicle 

29 microenvironments (U.S. EPA, 2014, section 5.1.3). APEX takes into account the most 

 

59In 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. 

60In 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. 
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1 significant factors that contribute to total human exposure to ambient O3, including the temporal 

2 and spatial distributions of people and O3 concentrations throughout an urban area, the variation 

3 of O3 concentrations within various microenvironments, and the effects of exertion on breathing 

4 rate in exposed individuals (U.S. EPA, 2014, section 5.1.3). To the extent spatial and/or temporal 

5 patterns of ambient O3 concentrations are altered upon model adjustment, as discussed above, 

6 exposure estimates reflect population exposures to those altered patterns. 

7 The second draft HREA estimates 8-hour exposures at or above benchmark 

8 concentrations of 60, 70, and 80 ppb for individuals engaged in moderate or greater exertion. 

9 Benchmarks reflect exposure concentrations at which O3-induced respiratory effects are known 

10 to occur in some healthy adults engaged in moderate, intermittent exertion, based on evidence 

11 from controlled human exposure studies (section 3.1.2.1 above and U.S. EPA, 2013, section 6.2). 

12 The amount of weight to place on the estimates of exposures at or above specific benchmark 

13 concentrations depends in part on the weight of the scientific evidence concerning health effects 

14 associated with O3 exposures at that concentration. It also depends on judgments about the 

15 importance, from a public health perspective, of the health effects that are known or can 

16 reasonably be inferred to occur as a result of exposures at benchmark concentrations (sections 

17 3.1.3, 3.1.5 above). 

18 As discussed in more detail above (section 3.1.2.1), the health evidence that supports 

19 evaluating exposures of concern at or above benchmark concentrations of 60, 70, and 80 ppb 

20 comes from a large body of controlled human exposure studies reporting a variety of respiratory 

21 effects in healthy adults. The lowest O3 exposure concentration for which controlled human 

22 exposure studies have reported respiratory effects in healthy adults is 60 ppb, with more 

23 evidence supporting this benchmark concentration in the current review than in the last review. 

24 In healthy adults, exposures to 60 ppb O3 have been reported to decrease lung function and to 

25 increase airway inflammation. Exposures of healthy adults to 70 ppb O3 have been reported to 

26 result in larger lung function decrements, compared to 60 ppb, as well as in increased respiratory 

27 symptoms. Exposures of healthy adults to 80 ppb O3 have been reported to result in larger lung 

28 function decrements than following exposures to 60 or 70 ppb, increased airway inflammation, 

29 increased respiratory symptoms, increased airways responsiveness, and decreased lung host 

30 defense (section 3.1.2.1, above). As discussed above (section 3.1.3), respiratory effects reported 

31 following exposures to O3 concentrations of 60, 70, or 80 ppb meet ATS criteria for adverse 

32 effects, result in effects judged important by CASAC in past reviews, and/or could contribute to 

33 the clearly adverse effects reported in epidemiologic studies evaluating broader populations. 

34 Compared to the healthy individuals included in the studies that provided the basis for the 

35 benchmarks, at-risk populations (e.g., asthmatics, children) are more likely to experience larger 

36 and/or more serious effects (e.g., U.S. EPA 2013, p. 6-21). 
3-89

1 In considering estimates of O3 exposures of concern at or above benchmarks of 60, 70, 

2 and 80 ppb, within the context of the adequacy of the current standard, we first address the 

3 following specific question: 

4 What are the nature and magnitude of the short-term O3 exposures of concern 

5 remaining upon adjustment of air quality to just meet the current O3 standard? 

6 In addressing this question, we focus on modeled exposures for school-age children (ages 

7 5-18) and asthmatic school-age children, two of the at-risk populations identified in the ISA

8 (section 3.1.5 above). The percentages of children estimated to experience exposures of concern 

9 are considerably larger than the percentages estimated for adult populations (i.e., approximately 

10 3-fold larger across cities) (U.S. EPA, 2014, Figures 5-5, 5-6, 5-7). The larger exposure 

11 estimates for children are due primarily to the larger percentage of children estimated to spend an 

12 extended period of time being physically active outdoors (U.S. EPA, 2014, section 5.4.1, Figure 

13 5-16). 

14 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),61 15 and we note that estimates for all 

16 children and asthmatic children are virtually indistinguishable (U.S. EPA, 2014, section 5.3.2). 

17 The estimates presented in Figures 3-7 to 3-10 below reflect consistent reductions in estimated 

18 exposures of concern across urban case study areas, relative to recent (i.e., unadjusted) air quality 

19 (U.S. EPA, 2014, Appendix to Chapter 5). When averaged over the years evaluated in the 

20 HREA, reductions of up to about 70% were estimated, compared to recent air quality. These 

21 reductions in estimated exposures of concern, relative to unadjusted air quality, reflect the 

22 consistent reductions in the highest ambient O3 concentrations upon model adjustment to just 

23 meet the current standard (section 3.2.1 above; U.S. EPA, 2014, Chapter 4). 

24 Although exposure estimates differ between children and adults, the patterns of results 

25 across the cities and years are similar among all of the populations evaluated (U.S. EPA, 2014, 

26 Figures 5-12 to 5-15). Therefore, while we highlight estimates in children, we also note that the 

27 patterns of exposures estimated for children represent the patterns estimated for adult asthmatics 

28 and older adults. 

 

61Figures 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. 
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1

2 Figure 3-7. Percent of children estimated to experience one or more exposures of concern at or above 60, 70, 80 ppb with air 

3 quality adjusted to just meet the current standard (averaged over 2006 to 2010). 

4
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1

2

3

4

5 Figure 3-8. Percent of children estimated to experience one or more exposures of concern at or above 60, 70, 80 ppb with air 

6 quality adjusted to just meet the current standard (worst-case year, 2006 to 2010). 

7
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1

2 Figure 3-9. Percent of children estimated to experience two or more exposures of concern at or above 60, 70, 80 ppb with air 

3 quality adjusted to just meet the current standard (Averaged over 2006 to 2010) 

4
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1

2

3 Figure 3-10. Percent of children estimated to experience two or more exposures of concern at or above 60, 70, 80 ppb with air 

4 quality adjusted to just meet the current standard (Worst-Case Year, 2006 to 2010) 

5
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1 Based on Figures 3-7 to 3-10 and the associated details described in the second draft 

2 HREA (U.S. EPA 2014, Chapter 5), we take note of the following with regard to exposures that 

3 are estimated to be allowed by the current standard: 

4 1. For exposures of concern at or above 60 ppb: 

5 a. On average over the years 2006 to 2010, the current standard is estimated to allow 

6 approximately 10 to 18% of children in urban case study areas to experience one or 

7 more exposures of concern at or above 60 ppb. Summing across urban case study 

8 areas, these percentages correspond to almost 2.5 million children experiencing 

9 approximately 4 million exposures of concern at or above 60 ppb during a single O3

10 season. Of these children, almost 250,000 are asthmatics. 

11

12 b. On average over the years 2006 to 2010, the current standard is estimated to allow 

13 approximately 3 to 8% of children in urban case study areas to experience two or 

14 more exposures of concern to O3 concentrations at or above 60 ppb. Summing across 

15 the urban case study areas, these percentages correspond to almost 900,000 children 

16 (including almost 90,000 asthmatic children) estimated to experience at least two O3

17 exposure concentrations at or above 60 ppb during a single O3 season. 

18

19 c. In the worst-case years (i.e., those with the largest exposure estimates), the current 

20 standard is estimated to allow approximately 10 to 25% of children to experience one 

21 or more exposures of concern at or above 60 ppb, and approximately 4 to 14% to 

22 experience two or more exposures of concern at or above 60 ppb. 

23

24 2. For exposures of concern at or above 70 ppb: 

25 a. On average over the years 2006 to 2010, the current standard is estimated to allow up 

26 to approximately 3% of children in urban case study areas to experience one or more 

27 exposures of concern at or above 70 ppb. Summing across urban case study areas, 

28 almost 400,000 children (including almost 40,000 asthmatic children) are estimated to 

29 experience O3 exposure concentrations at or above 70 ppb during a single O3 season. 

30

31 b. On average over the years 2006 to 2010, the current standard is estimated to allow 

32 less than 1% of children in urban case study areas to experience two or more 

33 exposures of concern to O3 concentrations at or above 70 ppb. 

34

35 c. In the worst-case years, the current standard is estimated to allow approximately 1 to 

36 8% of children to experience one or more exposures of concern at or above 70 ppb, 

37 and up to approximately 2% to experience two or more exposures of concern, at or 

38 above 70 ppb. 

39

40 3. For exposures of concern at or above 80 ppb: The current standard is estimated to allow 

41 about 1% or fewer children in urban case study areas to experience exposures of concern at 

42 or above 80 ppb, even in years with the highest exposure estimates. 
3-95

In further evaluating estimated exposures of concern from the 2nd 1 draft HREA, we next 

2 consider the following question: 

3 What are the important sources of uncertainty associated with exposure estimates?

4 Due to variability in responsiveness, only a subset of individuals who experience 

5 exposures at or above a benchmark concentration can be expected to experience health effects. 

6 Given the lack of sufficient exposure-response information for most of the health effects that 

7 informed benchmark concentrations, estimates of the number of people likely to experience 

8 exposures at or above benchmark concentrations generally cannot be translated into quantitative 

estimates of the number of people likely to experience specific health effects. 9

62 We view health10 relevant exposures as a continuum with greater confidence and less uncertainty about the 

11 existence of health effects at higher O3 exposure concentrations, and less confidence and greater 

12 uncertainty as one considers lower exposure concentrations. This view draws from the overall 

13 body of available health evidence, which indicates that as exposure concentrations increase the 

14 incidence, magnitude, and severity of effects increases. 

15 Though we have less confidence in the likelihood of adverse health effects as O3

16 exposure concentrations decrease, we also note that the controlled human exposure studies that 

17 provided the basis for health benchmark concentrations have not evaluated at-risk populations. 

18 Compared to the healthy individuals included in controlled human exposure studies, members of 

19 at-risk populations (e.g., asthmatics, children) could be more likely to experience adverse effects, 

20 could experience larger and/or more serious effects, and/or could experience effects following 

21 exposures to lower O3 concentrations. In considering estimated exposures of concern within the 

22 context of drawing conclusions on the adequacy of the current standard (section 3.4, below), we 

23 balance concerns about the potential for adverse health effects, including effects in at-risk 

24 populations, with our increasing uncertainty regarding the likelihood of such effects following 

25 exposures to lower O3 concentrations. 

26 Uncertainties associated with the APEX exposure modeling also have the potential to be 

27 important in our consideration of the adequacy of the current standard. For example, the HREA 

28 concludes that exposures of concern could be underestimated for some individuals who are 

29 frequently and routinely active outdoors during the warm season (U.S. EPA, section 5.5.2). This 

30 could include outdoor workers and children who are frequently active outdoors. The HREA 

31 specifically notes that long-term diary profiles (i.e., monthly, annual) do not exist for such 

32 populations, limiting the extent to which APEX outputs reflect people who follow similar daily 

33 routines resulting in high exposures, over extended periods of time. Thus, exposure estimates 

 

62The exception to this is lung function decrements, as discussed below (section 3.2.3.1). 
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1 generated from the general pool of available diary profiles likely do not reflect the most highly 

2 exposed individuals in the population. 

3 In order to evaluate the potential implications of this uncertainty for exposure estimates, 

4 the second draft HREA reports the results of limited sensitivity analyses using subsets of diaries 

5 specifically selected to reflect groups spending a larger proportion of time being active outdoors 

6 during the O3 season. When diaries were selected to mimic exposures that could be experienced 

7 by outdoor workers, the percent of modeled individuals estimated to experience exposures of 

8 concern increased compared to other adult populations evaluated. The percent of outdoor 

9 workers estimated to experience exposures of concern were generally similar to the percentages 

10 estimated for children (i.e., using the full database of diary profiles) in the worst-case cities and 

11 years (i.e., cities and years with the highest exposure estimates) (U.S. EPA, 2014, Figure 5-11). 

12 In addition, when diaries were restricted to children who did not report any time spent inside a 

13 school or performing paid work (i.e., to mimic children spending particularly large portions of 

14 their time outdoors during the summer), the number experiencing exposures of concern increased 

15 by approximately 30% (U.S. EPA, 2014, section 5.3.3). Though these sensitivity analyses are 

16 limited to single urban case study areas, and though there is uncertainty associated with diary 

17 selection approaches to mimic highly exposed populations, they suggest the possibility that some 

18 portions of the population could experience more frequent exposures of concern than indicated 

19 by estimates based on the full database of activity diary profiles. 

20 In further considering activity diaries, the HREA also notes that growing evidence 

21 indicates that people can change their behavior in response to high O3 concentrations, reducing 

22 the time spent being active outdoors (U.S. EPA, 2014, section 5.4.3). Commonly termed 

23 “averting behaviors,” these altered activity patterns could reduce personal exposure 

24 concentrations. Therefore, the second draft HREA also performed limited sensitivity analyses to 

25 evaluate the potential implications of averting behavior for estimated exposures of concern. 

26 These analyses suggest that averting behavior could reduce the percentages of children estimated 

27 to experience exposures of concern at or above the 60 or 70 ppb benchmark concentrations by 

28 approximately 10 to 30%, with larger reductions possible for the 80 ppb benchmark (U.S. EPA, 

29 2014, Figure 5-12). As discussed above for other sensitivity analyses, these analyses are limited 

30 to a single urban case study area and are subject to uncertainties associated with assumptions 

31 about the prevalence and duration of averting behaviors. However, the results suggest that 

32 exposures of concern could be overestimated, particularly in children (Neidell et al., 2009; U.S. 

33 EPA, 2013, Figures 4-7 and 4-8), if the possibility for averting behavior is not incorporated into 

34 estimates. 
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1 3.2.3 Risk-Based Considerations 

2 For some health endpoints, there is sufficient scientific evidence and information 

3 available to support the development of quantitative estimates of O3-related health risks. In the 

4 last review of the O3 NAAQS, the quantitative health risk assessment estimated O3-related lung 

5 function decrements, respiratory symptoms, respiratory-related hospital admissions, and non6 accidental and cardiorespiratory-related mortality (U.S. EPA, 2007). In those analyses, both 

7 controlled human exposure and epidemiologic studies were used for the quantitative assessment 

8 of O3-related human health risks. 

9 In the current review, for short-term O3 concentrations the second draft HREA estimates 

10 lung function decrements; respiratory symptoms in asthmatics; hospital admissions and 

11 emergency department visits for respiratory causes; and all-cause mortality (U.S. EPA, 2014). 

12 For “long-term” O3 concentrations, the second draft HREA estimates respiratory mortality (U.S. 

EPA, 2014).63 13 Estimates of O3-induced lung function decrements are based on exposure 

14 modeling, as noted above, combined with exposure-response relationships from controlled 

15 human exposure studies (U.S. EPA, 2014, Chapter 6). Estimates of O3-associated respiratory 

16 symptoms; hospital admissions and emergency department visits; and mortality are based on 

17 concentration-response relationships from epidemiologic studies (U.S. EPA, 2014, Chapter 7). 

18 As with the exposure assessment discussed above, O3-associated health risks are estimated for 

19 recent air quality and for ambient concentrations model-adjusted to just meet the current 8-hour 

20 O3 NAAQS, based on 2006-2010 air quality and adjusted precursor emissions. 

21 Section 3.2.3.1 below discusses risk results for O3-induced lung function decrements 

22 following short-term exposures, based on exposure modeling results and exposure-response 

23 relationships from controlled human exposure studies. Section 3.2.3.2 discusses epidemiology24 based risk estimates, with a focus on all-cause mortality (short-term O3 concentrations); 

25 respiratory-related morbidity outcomes (short-term O3 concentrations); and respiratory mortality 

26 (long-term O3 concentrations). 

27 3.2.3.1 Risk of Lung Function Decrements 

28 In the last review, EPA conducted a health risk assessment that produced risk estimates 

29 for the number and percent of school-aged children, asthmatic school-aged children, and the 

30 general population experiencing lung function decrements associated with O3 exposures for 12 

31 urban areas. These estimates were based on exposure-response relationships developed from 

 

63 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. 
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1 analysis of data from several controlled human exposure studies, combined with exposure 

2 estimates developed for children and adults (U.S. EPA, 2007a). 

3 In the current review, the second draft HREA estimates risks of lung function decrements 

4 in school-aged children (ages 5 to 18), asthmatic school-aged children, and the general adult 

population for 15 urban case study areas.64 5 The results presented in the second draft HREA are 

6 based on an updated dose-threshold model that estimates FEV1 responses for individuals 

7 following short-term exposures to O3 (McDonnell, Stewart, and Smith, 2010), reflecting 

8 methodological improvements since the last review (U.S. EPA, 2014, section 6.2.4). The impact 

9 of the dose threshold is that O3-induced FEV1 decrements result primarily from exposures on 

10 days with ambient O3 concentrations above about 40 ppb (U.S. EPA, 2013, Chapter 6). 

11 As discussed above (section 3.1.3), for effects such as lung function decrements, which 

12 are transient and reversible, aspects such as the likelihood that these effects would occur 

13 repeatedly and would interfere with normal activities are important to consider in making 

14 judgments about adversity to individuals. As stated in the 2006 Criteria Document (Table 8-3, 

15 p.8-68), for people with lung disease even moderate functional responses (e.g., FEV1 decrements 

16 ≥ 10% but < 20%) would likely interfere with normal activities for many individuals, and would 

17 likely result in more frequent medication use. Moreover, as noted above, in a recent letter to the 

18 Administrator, the CASAC O3 Panel stated that “'[c]linically relevant effects are decrements > 

19 10%, a decrease in lung function considered clinically relevant by the American Thoracic 

20 Society” (Samet, 2011, p.2). The CASAC O3 Panel also stated that: 

21 [A] 10% decrement in FEV1 can lead to respiratory symptoms, especially in 

22 individuals with pre-existing pulmonary or cardiac disease. For example, people 

23 with chronic obstructive pulmonary disease have decreased ventilatory reserve 

24 (i.e., decreased baseline FEV1) such that a ≥ 10% decrement could lead to 

25 moderate to severe respiratory symptoms (Samet, 2011, p.7).

26 In judging the extent to which moderate lung function decrements represent effects that 

27 should be regarded as adverse to the health status of individuals, in previous NAAQS reviews we 

28 have also considered the extent to which decrements were experienced repeatedly during the 

29 course of a year (Staff Paper, U.S. EPA, 2007). Although some experts would judge single 

30 occurrences of moderate responses to be a ‘‘nuisance,’’ especially for healthy individuals, a 

31 more general consensus view of the adversity of such moderate responses emerges as the 

32 frequency of occurrence increases. Thus in the past EPA has judged that repeated occurrences of 

33 moderate responses, even in otherwise healthy individuals, may be considered to be adverse 

 

64As 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. 
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1 since they could well set the stage for more serious illness (61 FR 65723). The CASAC panel in 

2 the 1997 NAAQS review expressed a consensus view that these “criteria for the determination of 

3 an adverse physiological response were reasonable” (Wolff, 1995). In the review completed in 

4 2008, estimates of repeated occurrences continued to be an important public health policy factor 

5 in judging the adversity of moderate lung function decrements in healthy and asthmatic 

6 populations (72 FR 37850, July 11, 2007). 

7 The second draft HREA estimates risks of moderate to large lung function decrements, 

8 defined as FEV1 decrements > 10%, > 15%, or > 20%. In evaluating these lung function risk 

9 estimates within the context of considering the adequacy of the current O3 standard, we first 

10 consider the following specific question: 

11 What are the nature and magnitude of lung function risks remaining upon just meeting 

12 the current O3 standard? 

13 In considering risks of O3-induced FEV1 decrements, we focus on the percent of children 

14 estimated to experience decrements > 10, 15, and 20%, noting that the percentage of asthmatic 

15 children estimated to experience such decrements is virtually the same as the percentage 

16 estimated for all children. Compared to children, only a very small percentage of adults were 

17 estimated to experience O3-induced FEV1 decrements (U.S. EPA, 2014, Appendix 6B). As for 

18 exposures of concern (see above), the patterns of results across urban case study areas and over 

19 the years evaluated are similar in children and adults (U.S. EPA, 2014, Chapter 6). Therefore, 

20 while we highlight estimates in children, we note that these results are also representative of the 

21 patterns estimated for adult populations. 

22 Key results for children are summarized below for air quality adjusted to just meet the 

current O3 NAAQS (Figures 3-11 to 3-14).65 23 The estimates presented in Figures 3-11 to 3-14 

24 below reflect consistent reductions across urban case study areas in the percent of children 

25 estimated to experience O3-induced lung function decrements, relative to recent (i.e., unadjusted) 

26 air quality (U.S. EPA, 2014, Appendix to Chapter 6). When averaged over the years evaluated 

27 in the HREA, reductions of up to about 40% were estimated compared to recent air quality. 

28 These reductions reflect the consistent decreases in relatively high ambient O3 concentrations 

29 upon adjustment to just meet the current standard (section 3.2.1 above; U.S. EPA, 2014, Chapter 

4).66 30 

 

65Figures 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. 

66As noted above, the impact of the dose threshold is that O3-induced FEV1 decrements result primarily from days 

with ambient O3 concentrations above about 40 ppb (U.S. EPA, 2013, Chapter 6). 
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1

2 Figure 3-11. Percent of school-aged children (5-18 yrs) estimated to experience one or more days with FEV1 decrements > 10, 

3 15, or 20% with air quality adjusted to just meet the current standard (Averaged over 2006 to 2010) 

4
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1

2

3 Figure 3-12. Percent of school-aged children (5-18 yrs) estimated to experience one or more days with FEV1 decrements > 10, 

4 15, or 20% with air quality adjusted to just meet the current standard (Worst-Case Year from 2006 to 2010) 

5
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1 

2 Figure 3-13. Percent of school-aged children (5-18 yrs) estimated to experience two or more days with FEV1 decrements > 10, 

3 15, or 20% with air quality adjusted to just meet the current standard (Averaged over 2006 to 2010) 

4
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1

2 Figure 3-14. Percent of school-aged children (aged 5-18 yrs) estimated to experience two or more days with FEV1 decrements >

3 10, 15, or 20% with air quality adjusted to just meet the current standard (Worst-Case Year from 2006 to 2010)
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1 Based on Figures 3-11 to 3-14 and the associated details described in the second draft 

2 HREA (U.S. EPA 2014, Chapter 6), we take note of the following with regard to lung function 

3 decrements estimated to be allowed by the current standard: 

4 1. For FEV1 decrements > 10%: 

5 a. On average over the years 2006 to 2010, the current standard is estimated to allow 

6 approximately 14 to 19% of children in urban case study areas to experience one or 

7 more lung function decrements > 10%. Summing across urban case study areas, this 

8 corresponds to approximately 3 million children experiencing 15 million O3-induced 

9 lung function decrements > 10% during a single O3 season. Of these children, about 

10 300,000 are asthmatics. 

11

12 b. On average over the years 2006 to 2010, the current standard is estimated to allow 

13 approximately 7 to 12% of children in urban case study areas to experience two or 

14 more O3-induced lung function decrements > 10%. Summing across the urban case 

15 study areas, this corresponds to almost 2 million children (including almost 200,000 

16 asthmatic children) estimated to experience two or more O3-induced lung function 

17 decrements greater than 10% during a single O3 season. 

18

19 c. In the worst-case years, the current standard is estimated to allow approximately 17 to 

20 23% of children in urban case study areas to experience one or more lung function 

21 decrements > 10%, and approximately 10 to 14% to experience two or more O3-

22 induced lung function decrements > 10%. 

23

24 2. For FEV1 decrements > 15%: 

25 a. On average over the years 2006 to 2010, the current standard is estimated to allow 

26 approximately 3 to 5% of children in urban case study areas to experience one or 

27 more lung function decrements > 15%. Summing across urban case study areas, this 

28 corresponds to approximately 800,000 children (including approximately 80,000 

29 asthmatic children) estimated to experience at least one O3-induced lung function 

30 decrement > 15% during a single O3 season. 

31

32 b. On average over the years 2006 to 2010, the current standard is estimated to allow 

33 approximately 2 to 3% of children in urban case study areas to experience two or 

34 more O3-induced lung function decrements > 15%. 

35

36 c. In the worst-case years, the current standard is estimated to allow approximately 4 to 

37 6% of children in urban case study areas to experience one or more lung function 

38 decrements > 15%, and approximately 2 to 4% to experience two or more O3-induced 

39 lung function decrements > 15%. 

40

41 3. For FEV1 decrements > 20%: 

42 a. On average over the years 2006 to 2010, the current standard is estimated to allow 

43 approximately 1 to 2% of children in urban case study areas to experience one or 

44 more lung function decrements > 20%. Summing across urban case study areas, this 

45 corresponds to approximately 300,000 children (including approximately 30,000 
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1 asthmatic children) estimated to experience at least one O3-induced lung function 

2 decrement > 20% during a single O3 season. 

3

4 b. On average over the years 2006 to 2010, the current standard is estimated to allow 

5 less than 1% of children in urban case study areas to experience two or more O3-

6 induced lung function decrements > 20%. 

7

8 c. In the worst-case years, the current standard is estimated to allow approximately 2 to 

9 3% of children to experience one or more lung function decrements > 20%, and less 

10 than 2% to experience two or more O3-induced lung function decrements > 20%.

In further considering estimated lung function risks from the 2nd 11 draft HREA, we next 

12 consider the following question: 

13 What are the important sources of uncertainty associated with lung function risk 

14 estimates?

15 In addition to the uncertainties noted above for exposure estimates, the HREA identifies 

16 several key uncertainties associated with estimates of O3-induced lung function decrements. An 

17 uncertainty with particular potential to impact our consideration of risk estimates in this Policy 

18 Assessment stems from the lack of exposure-response information in children. In the absence of 

19 controlled human exposure data for children, risk estimates are based on the assumption that 

20 children exhibit the same lung function response following O3 exposures as healthy 18 year olds 

21 (i.e., the youngest age for which controlled human exposure data is available) (U.S. EPA, 2014, 

22 section 6.2.4 and 6.5). This assumption was justified in part by the findings of McDonnell et al. 

23 (1985), who reported that children 8-11 year old experienced FEV1 responses similar to those 

24 observed in adults 18-35 years old. In addition, as discussed in the ISA (U.S. EPA, 2013, section 

25 6.2.1), summer camp studies of school-aged children reported O3-induced lung function 

26 decrements similar in magnitude to those observed in controlled human exposure studies using 

27 adults. In extending the risk model to children, the second draft HREA fixes the age term in the 

28 model at its highest value, the value for age 18. This approach could result in either over- or 

29 underestimates of O3-induced lung function decrements in children, depending on how children 

30 compare to the adults used in controlled human exposure studies (U.S. EPA, 2014, section 6.5). 

31 A related source of uncertainty is that the risk assessment estimates O3-induced 

32 decrements in asthmatics using the exposure-response relationship developed from data collected 

33 from healthy individuals. Although the evidence has been mixed (U.S. EPA, 2013, section 

34 6.2.1.1), several studies have reported larger O3-induced lung function decrements in asthmatics 

35 than in non-asthmatics (Kreit et al., 1989; Horstman et al., 1995; Jorres et al., 1996; Alexis et al., 

36 2000). To the extent asthmatics experience larger O3-induced lung function decrements than the 

37 healthy adults used to develop exposure-response relationships, the second draft HREA could 
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1 underestimate the impacts of O3 exposures on lung function in asthmatics, including asthmatic 

2 children. The HREA notes that the magnitude this uncertainty might have on risk estimates 

3 remains unknown at this time (U.S. EPA, 2014, Chapter 6). 

4 3.2.3.2 Estimated Health Risks Associated with Short- or Long-Term O3 Exposures, 

5 Based on Epidemiologic Studies 

6 Risk estimates based on epidemiologic studies can provide perspective on the most 

7 serious O3-associated public health outcomes (e.g., mortality, hospital admissions, emergency 

8 department visits) in populations that include at-risk groups. The second draft HREA estimates 

O3-associated risks in 12 urban case study areas67 9 using concentration-response relationships 

10 drawn from epidemiologic studies. These concentration-response relationships are based on 

“area-wide” average O3 concentrations.68 11 The HREA estimates risks for the years 2007 and 2009 

12 in order to provide estimates of risk for a year with generally higher O3 concentrations (2007) 

13 and a year with generally lower O3 concentrations (2009) (U.S. EPA, 2014, section 7.2). 

14 In the last review, epidemiologic-based risks were estimated for O3 concentrations above 

15 mean “policy-relevant background concentrations.” As discussed above (Chapter 2), policy16 relevant background (PRB) concentrations were defined as the distribution of ozone 

17 concentrations that would be observed in the U.S. in the absence of anthropogenic (man-made) 

18 emissions of ozone precursor emissions (e.g., VOC, CO, NOx) in the U.S., Canada, and Mexico. 

19 This approach provided a focus on O3 concentrations “that can be controlled by U.S. regulations 

20 (or through international agreements with neighboring countries)” (U.S. EPA, 2007, pp. 2-48 to 

21 2-54). 

22 As in the last review, we recognize that ambient O3 concentrations, and therefore O3-

23 associated health risks, result from precursor emissions from various types of sources. Based on 

24 the air quality modeling discussed above in chapter 2, approximately 30 to 60% of average 

25 daytime O3 during the warm season (i.e., 8-hour daily maximum concentrations averaged from 

26 April to October) is attributable to precursor emissions from U.S. anthropogenic sources (section 

27 2.4.4). This suggests that, for recent air quality (i.e., not adjusted to meet the current or 

28 alternative standards), approximately 30 to 60% of total O3-associated health risk in the urban 

 

67 The 12 urban areas evaluated are Atlanta, Baltimore, Boston, Cleveland, Denver, Detroit, Houston, Los Angeles, 

New York, Philadelphia, Sacramento, and St. Louis. 

68In 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 second draft 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. 
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1 case study areas is attributable to O3 from U.S. anthropogenic emissions. The remainder is 

2 attributable to precursor emissions from international anthropogenic sources and natural sources. 

3 Because the second draft HREA characterizes health risks from all O3, regardless of source, risk 

4 estimates reflect emissions from U.S. anthropogenic, international anthropogenic, and natural 

5 sources. Given that HREA risk estimates for adjusted air quality are based on decreasing U.S. 

6 anthropogenic precursor emissions, the contributions of U.S. anthropogenic emissions to the risk 

7 estimates for the current standard would generally be smaller than the 30 to 60% indicated for 

8 recent air quality. 

9 In evaluating epidemiology-based risk estimates within the context of the adequacy of the 

10 current standard, we first consider the following question: 

11 What are the nature and magnitude of the O3-associated mortality and morbidity risks 

12 remaining upon adjustment of air quality to just meet the current O3 standard? 

13 In addressing this question, as an initial matter we note that the area-wide average O3

14 concentrations associated with health effects in epidemiologic studies, and used to estimate 

15 mortality and morbidity risks in the HREA, are surrogates for the ambient O3 exposures expected 

16 to have elicited the reported health outcomes (also discussed in section 3.1.4.2, above). The area17 wide average concentrations present in epidemiologic study locations represent the spatial and 

18 temporal patterns of O3 exposures (magnitudes, frequencies, durations of exposures) experienced 

19 by study populations. Differences in the patterns of O3 exposures would be expected to result in 

20 differences in the health outcome response. Thus, in considering the quantitative risk estimates 

21 below we are mindful of uncertainties related to the differences between the spatial and temporal 

22 patterns of O3 that existed in the epidemiologic study areas, which contributed to the health 

23 outcomes reported in these studies, and the altered patterns associated with adjusted air quality 

24 that just meets the current standard (section 3.2.1, above). Among the three main types of 

25 exposure/risk analyses generated in the HREA, these altered spatial/temporal patterns have the 

26 greatest potential to introduce uncertainty into risk estimates based on epidemiology study 

concentration-response relationships (see 2nd 27 question below for further discussion). 

28 We also note that the second draft HREA estimates mortality and morbidity risks 

29 associated with just meeting the current standard by applying concentration-response 

30 relationships from epidemiologic studies to the entire distributions of model-adjusted “area31 wide” average O3 concentrations present in urban case study areas (U.S. EPA, 2014, Chapter 7). 

32 Implicit in this approach to estimating risks is the assumption that concentration-response 

33 relationships are linear over those distributions. Therefore, as noted in section 3.2.1, when air 

34 quality is adjusted to just meet the current standard, risk estimates are influenced by the 

35 decreases in area-wide O3 concentrations at the upper ends of warm season distributions and the 
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1 increases in area-wide O3 concentrations at the lower ends of those distributions (U.S. EPA, 

2014, section 4.3.3.2, Figures 4-8 to 4-11). 2

69 When the decreases and increases are of the same 

3 magnitude, they result in the same degree of change in estimated risks, though opposite in 

4 direction. Therefore, seasonal estimates of O3-associated mortality and morbidity risks either 

5 increase or decrease in response to air quality adjustment, depending on the seasonal balance 

6 between the modeled daily decreases in high area-wide O3 concentrations and increases in low 

7 area-wide O3 concentrations. One consequence is that the estimated impacts on mortality and 

8 morbidity risks of adjusting air quality to just meet the current standard are more modest, and 

9 less directionally consistent across urban case study areas, than on either exposures of concern or 

10 O3-induced lung function decrements. 

11 In the remainder of this section, we consider estimates of total (non-accidental) mortality 

12 and respiratory morbidity associated with short-term O3 concentrations, and respiratory mortality 

13 associated with “long-term” O3 concentrations. 

14 Total Mortality – Short-Term O3

15 Risk estimates for total mortality are based on concentration-response relationships 

16 described by Smith et al. (2009). To generate risk estimates, the second draft HREA uses “area17 wide” averages of 8-hour daily maximum O3 concentrations over the full monitoring periods in 

18 urban case study areas. These monitoring periods vary across areas, in some cases including 

19 more of the cooler months that are often characterized by relatively low daytime O3

concentrations. 20 70 When air quality was adjusted to the current standard in the 2007 model year 

21 (the year with generally “higher” O3-associated risks), 10 of 12 urban case study areas exhibited 

22 either small decreases or virtually no change in estimates of O3-associated total mortality (U.S. 

EPA, 2014, Appendix to Chapter 7).71 23 Small increases in mortality were estimated in two of the 

24 urban case study areas (Houston, Los Angeles) (U.S. EPA, 2014, Appendix to Chapter 7). 

25 Figure 3-15 below presents estimates of O3-associated all-cause mortality in urban case 

26 study areas for 2007 and 2009, with air quality adjusted to just meet the current O3 standard. The 

27 second draft HREA estimates that upon just meeting the current standard, O3 could be associated 

28 with from 0.8 to 4.1% of all-cause mortality across the urban case study areas. This corresponds 

 

69On a given day, area-wide O3 concentrations and estimated risks decrease when the sum of the changes at monitors 

with decreasing O3 are larger than the sum of the changes at monitors with increasing O3. Area-wide O3

concentrations and estimated risks increase when the opposite occurs. 

70Decreases in relatively higher ambient O3 concentrations are more prominent during the warmest months, when 

daytime concentrations tend to be highest. In most urban case study areas, increases in relatively low daytime 

concentrations have greater influence on risk estimates during cooler months, when O3 concentrations tend to be 

lower overall (U.S. EPA, 2014, compare Figures 4-8 and 4-9). 

71Decreases were smaller in the 2009 model-adjusted year (i.e., the year with generally lower O3 concentrations). 
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1 to approximately 80 to 2,800 O3-associated deaths per season in individual urban case study 

2 areas, and approximately 8,000 to 9,000 O3-associated deaths per season summed over the 12 

3 urban case study areas (U.S. EPA, 2014, Tables 7-7 and 7-8). 

4

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

6 just meet the current standard. 

7 In considering the risk estimates presented in Figure 3-15, which are based on applying 

8 linear concentration-response relationships to the full distributions of daily 8-hour “area-wide” 

9 O3 concentrations, we note the ISA conclusion that there is less certainty in the shape of 

10 concentration-response functions for area-wide O3 concentrations at the lower ends of warm 

11 season distributions (i.e., below about 20 to 40 ppb depending on the O3 metric, health endpoint, 

12 and study population) (U.S. EPA, 2013, section 2.5.4.4). We also recognize that for the range of 

13 health endpoints evaluated, controlled human exposure and animal toxicological studies provide 

14 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).72 15 Thus, in addition to 

16 considering estimates of total O3-associated risks, we also consider the extent to which risks are 

17 associated with days with higher, versus lower, area-wide O3 concentrations. 

18 Figure 3-16 presents risk estimates, summed across urban case study areas, for days with 

19 area-wide concentrations at or above 20, 40, and 60 ppb. Daytime O3 concentrations in the upper 

 

72As 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. 
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1 portion of the distribution of area-wide concentrations (e.g., at or above 40 or 60 ppb) are 

2 estimated to be associated with hundreds to thousands of deaths per year in urban case study 

areas.73 3

4

5 Figure 3-16. Estimated O3-associated mortality attributable to days above various area6 wide average O3 concentrations, with air quality adjusted to just meet current 

7 standard (2007 Model Adjustment Year). 

8 Respiratory Mortality – “Long-Term” O3 

9 The second draft HREA estimates the risk of respiratory mortality associated with long10 term O3 exposures, based on the study by Jerrett et al. (2009) (U.S. EPA, 2014, Chapter 7). To 

11 generate risk estimates, the second draft HREA uses “area-wide” averages of 1-hour daily 

12 maximum O3 concentrations during the warm season (April to September). When air quality was 

13 adjusted to just meet the current standard 11 of the 12 urban case study areas exhibited modest 

14 decreases in estimated O3-associated respiratory mortality (i.e., compared to recent, unadjusted 

15 air quality). Risk estimates remained virtually unchanged in the remaining urban case study area 

16 (i.e., Los Angeles). Risk estimates for air quality adjusted to just meet the current standard are 

presented below in Figure 3-17.

74 17

18

 

73The 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. 

74The second draft HREA does not characterize distributions of respiratory mortality risks over distributions of 

ambient O3 concentrations. Therefore, in considering respiratory mortality risks we evaluate only estimates of total 

risk. 
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1

2 Figure 3-17. Percent of baseline respiratory mortality estimated to be associated with long3 term O3. 

4 Across urban case study areas, O3 is estimated to be associated with approximately 16 to 

5 20% of respiratory mortality during the warm season. This corresponds to approximately 500 to 

6 2,800 O3-associated deaths per season across areas, and a total of approximately 12,000 O3-

7 associated deaths in all 12 urban case study areas. 

8 Hospital Admissions, Emergency Department Visits, and Asthma Exacerbations 

9 Risk estimates for respiratory-related hospital admissions, emergency department visits, 

10 and asthma exacerbations associated with air quality adjusted to just meet the current standard 

11 are based on several studies, as presented in Table 7-2 of the second draft HREA (U.S. EPA, 

2014).75 12 Estimates indicate that O3-associated respiratory-related hospital admissions account for 

13 approximately 2 to 3% of total respiratory-related admissions in urban case study locations. 

14 Depending on the city, this corresponds to 10’s to 100’s of O3-associated hospital admissions per 

15 season. Estimates indicate that O3-associated respiratory-related emergency department visits 

16 account for approximately 3 to 20% of total respiratory-related emergency department visits in 

 

75As with respiratory mortality above, the second draft 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. 
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1 Atlanta (approximately 4,000 to 8,000 visits per season), and that O3-associated asthma 

2 exacerbations account for approximately 15 to 30% of total exacerbations in Boston (45,000 to 

3 130,000 exacerbations per season). Full estimates are presented in Tables 7-9 to 7-11 in the 

4 second draft HREA (U.S. EPA, 2014). 

5 Based on Figures 3-15 to 3-17 above, and the more detailed information presented in 

6 Chapter 7 of the second draft HREA (U.S. EPA, 2014), we note the following key observations: 

7 1. In focusing on total risk, the current standard is estimated to allow thousands of O3-

8 associated deaths per year in the urban case study areas. These estimates are based on 

9 concentration-response functions from epidemiologic studies that used either 8-hour daily 

10 O3 concentrations (total mortality associated with short-term O3) or seasonal averages of 

11 1-hour daily O3 concentrations (respiratory mortality associated with long-term O3). 

12

13 2. In focusing on the risks associated with the upper portions of distributions of ambient 

14 concentrations, the current standard is estimated to allow hundreds to thousands of O3-

15 associated deaths per year in the urban case study areas. These estimates are based on 

16 concentration-response functions from an epidemiologic study that evaluated associations 

17 between 8-hour daily O3 concentrations and total mortality. 

18

19 3. In urban case study areas, the current standard is estimated to allow tens to thousands of 

20 O3-associated morbidity events per year. Distributions of O3-associated morbidity over 

21 distributions of ambient O3 concentrations would likely be similar to mortality, though 

22 the second draft HREA did not analyze such distributions for morbidity endpoints. 

In further considering estimated O3-associated mortality and morbidity risks from the 2nd 23

24 draft HREA, we next consider the following question: 

25 What are the important sources of uncertainty associated with mortality and morbidity 

26 risk estimates?

27 Upon adjusting air quality to the current standard, O3-associated mortality and morbidity 

28 risks generally decrease in locations and time periods with relatively high ambient O3

29 concentrations and increase in locations and time periods with relatively low concentrations. 

30 Therefore, an important consideration for epidemiology-based risk estimates is the extent to 

31 which seasonal risk estimates in urban case study areas represent the U.S. as a whole, in terms of 

32 the O3 response to decreasing precursor emissions. To address this, the second draft HREA 

33 conducted national air quality analyses evaluating the response of ambient O3 concentrations to 

34 reductions in NOX emissions. Those analyses indicate that the 12 urban case study areas may not 

35 represent the response of O3 in other populated areas of the U.S., including suburban areas, 

36 smaller urban areas, and rural areas, and that the majority of the U.S. population lives in 

37 locations where reducing NOX emissions would be expected to result in decreases in warm 
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1 season averages of daily maximum 8-hour ambient O3 concentrations. One implication of this is 

2 that HREA estimates for the urban case study areas are likely to understate the average reduction 

3 in O3-assocaiated mortality and morbidity risk that would be experienced across the population 

4 upon reducing NOX emissions (U.S. EPA, 2014, Chapter 8). 

5 Section 7.4 of the second draft HREA also highlights some additional uncertainties 

6 associated with epidemiologic-based risk estimates (U.S. EPA, 2014). This section of the HREA 

7 identifies and discusses sources of uncertainty and presents a qualitative evaluation of key 

8 parameters that can introduce uncertainty into risk estimates (U.S. EPA, 2014, Table 7-4). For 

9 several of these parameters the HREA also presents quantitative sensitivity analyses (U.S. EPA, 

10 2014, sections 7.4.2 and 7.5.3). Of the uncertainties discussed in Chapter 7 of the HREA, those 

11 related to the application of concentration-response functions from epidemiologic studies can 

12 have particularly important implications for our consideration of epidemiology-based risk 

13 estimates in this second draft PA. 

14 As noted above, an important uncertainty is the shape of concentration-response 

15 functions at low ambient O3 concentrations (U.S. EPA, 2014, Table 7-4). Consistent with the 

16 ISA conclusion that there is no discernible population threshold in O3-associated health effects, 

17 the second draft HREA estimates epidemiology-based mortality and morbidity risks for entire 

18 distributions of ambient O3 concentrations, with the assumption that concentration-response 

19 relationships remain linear over those distributions. In addition, in recognition of the ISA 

20 conclusion that certainty in the shape of O3 concentration-response functions decreases at low 

21 ambient concentrations, the second draft HREA also estimates distributions of total mortality 

22 incidence for various portions of the distribution of ambient O3 concentrations. In this second 

23 draft PA, we consider both types of risk estimates while recognizing that we have greater 

24 certainty in the increased incidence and severity of O3-attributable effects at higher ambient O3

25 concentrations (which drive higher exposure concentrations, section 3.2.2 above), as compared 

26 to lower concentrations. 

27 The second draft HREA also notes important uncertainties associated with using a 

28 concentration-response relationship developed for a particular population in a particular location 

29 to estimate health risks in different populations and locations (U.S. EPA, 2014, Table 7-4). As 

30 discussed above, concentration-response relationships derived from epidemiologic studies reflect 

31 the spatial and temporal patterns of population exposures during the study. The second draft 

32 HREA applies concentration-response relationships from epidemiologic studies to adjusted air 

33 quality in study areas that are different from, and often larger in spatial extent than, the areas 

34 used to generate the relationships. This approach ensures the inclusion of the actual non35 attainment monitors that often determine the magnitude of emissions reductions for the air 

36 quality adjustments throughout the urban case study areas. This approach also allows the HREA 
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1 to estimate patterns of health risks more broadly across a larger area, including a broader range 

2 of air quality concentrations and a larger population. The second draft HREA notes that it is not 

3 possible to quantify the impacts of this uncertainty on risk estimates in most urban case study 

4 locations, though the HREA notes that mortality effect estimates for different portions of the 

5 New York City CBSA-based assessment area vary by a factor of almost 10 (U.S. EPA, 2014, 

6 section 7.5.3). 

7 An additional, related uncertainty is that associated with applying concentration-response 

8 functions from epidemiologic studies to adjusted air quality. Concentration-response functions 

9 from the O3 epidemiologic studies used in the HREA are based on associations between day to 

10 day variation in “area-wide” O3 concentrations (i.e., averaged across multiple monitors) and 

11 variation in health effects. Epidemiologic studies use these area-wide O3 concentrations, which 

12 reflect the particular spatial and temporal patterns of ambient O3 present in study locations, as 

13 surrogates for the pattern of O3 exposures experienced by study populations. To the extent 

14 adjusting O3 concentrations to just meet the current standard results in important alterations in 

15 the spatial and/or temporal patterns of ambient O3, there is uncertainty in the appropriateness of 

16 applying concentration-response functions from epidemiologic studies to estimate health risks 

associated with adjusted O3 air quality.76 17 Although the impact of this uncertainty on risk 

18 estimates cannot be quantified (U.S. EPA, 2014, Table 7-4), it has the potential to become more 

19 important as model adjustment results in larger changes in spatial and temporal patterns of 

20 ambient O3 concentrations across urban case study areas. 

21 There is also uncertainty related specifically to the public health importance of the 

22 increases in relatively low O3 concentrations following air quality adjustment. This uncertainty 

23 relates to the fact that risk estimates are equally influenced by decreasing high concentrations 

24 and increasing low concentrations, when the increases and decreases are of equal magnitude. 

25 Even on days with increases in relatively low area-wide average concentrations, resulting in 

26 increases in estimated risks, some portions of the urban case study areas could experience 

 

76As 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. 

Specifically, the relatively high O3 concentrations that often occur downwind of important NOX sources (e.g., 

outside urban centers) generally decrease, while the relatively low O3 concentrations near important NOX sources 

(e.g., in urban centers) generally increase (U.S. EPA, 2014, section 4.3.3.2). In addition, decreases or increases in 

ambient O3 could occur more broadly across areas in some instances, depending in part on meteorological 

conditions. Such decreases in high O3 concentrations and increases in low concentrations can result in compression 

of the spatial distributions of ambient O3 used to calculate area-wide average concentrations. Decreases and 

increases can also result in compression of the temporal distributions of the area-wide O3 concentrations used to 

estimate mortality and morbidity risks over a season, such that area-wide concentrations decrease on “high”-O3 days 

and increase on “low”-O3 days. As indicated in the second draft HREA (U.S. EPA, 2014, Figures 4-10 and 4-11), 

this compression of seasonal distributions of O3 concentrations is evident in all of the urban case study areas 

evaluated, though the degree of compression varies considerably across areas. The most dramatic compression 

occurs in Los Angeles (U.S. EPA, 2014, Figures 4-10 and 4-11). 
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1 decreases in high O3 concentrations. To the extent O3-attributable effects are more strongly 

2 supported for higher ambient concentrations, likely resulting in higher exposure concentrations 

3 for some portions of study areas, the impacts on risk estimates of increasing low O3

4 concentrations reflect an important source of uncertainty. 

5 Finally, we note the second draft HREA does not quantify any reductions in risk that 

6 could be associated with reductions in the ambient concentrations of pollutants other than O3, 

7 resulting from control of NOX. For example as discussed in chapter 2 of this second draft PA, 

8 NOx emissions contribute to ambient NO2, and NOx and VOCs can contribute to secondary 

9 formation of PM2.5 constituents, including ammonium sulfate (NH4SO4), ammonium nitrate 

10 (NH4NO3), and organic carbon (OC). Therefore, at some times and in some locations, control 

11 strategies that would reduce NOX emissions (i.e., to meet an O3 standard) could reduce ambient 

12 concentrations of NO2 and PM2.5, resulting in health benefits beyond those directly associated 

with reducing ambient O3 concentrations.77 13

14 3.3 CASAC ADVICE 

15 Following the 2008 decision to revise the primary O3 standard by setting the level at 

16 0.075 ppm (75 ppb), CASAC strongly questioned whether the standard met the requirements of 

17 the CAA, further described below. In September 2009, EPA announced its intention to 

18 reconsider the 2008 standards, issuing a notice of proposed rulemaking in January 2010 (FR 75 

19 2938). Soon after, EPA solicited CASAC review of that proposed rule and in January 2011 

20 solicited additional advice. This proposal was based on the scientific and technical record from 

21 the 2008 rulemaking, including public comments and CASAC advice and recommendations. As 

22 further described in section 1.2.2 above, EPA in the fall of 2011 did not revise the standard as 

23 part of the reconsideration process but decided to coordinate further proceedings on the 

24 reconsideration rulemaking with this ongoing periodic review. Accordingly, in this section we 

25 describe CASAC’s advice related to the 2008 final decision and the subsequent reconsideration, 

26 as well as its advice on the NAAQS review that was initiated in September 2008. 

27 In April 2008, the members of the CASAC Ozone Review Panel sent a letter to EPA 

28 stating “[I]n our most-recent letters to you on this subject—dated October 2006 and March 

29 2007—the CASAC unanimously recommended selection of an 8-hour average Ozone NAAQS 

30 within the range of 0.060 to 0.070 parts per million [60 to 70 ppb] for the primary (human 

31 health-based) Ozone NAAQS” (Henderson, 2008). The letter continued: 

32 The CASAC now wishes to convey, by means of this letter, its additional, 

33 unsolicited advice with regard to the primary and secondary Ozone NAAQS. In 

 

77We 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. 
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1 doing so, the participating members of the CASAC Ozone Review Panel are 

2 unanimous in strongly urging you or your successor as EPA Administrator to 

3 ensure that these recommendations be considered during the next review cycle for 

4 the Ozone NAAQS that will begin next year ... numerous medical organizations 

5 and public health groups have also expressed their support of these CASAC 

6 recommendations’ ... [The CASAC did] not endorse the new primary ozone 

7 standard as being sufficiently protective of public health. The CASAC—as the 

8 Agency’s statutorily-established science advisory committee for advising you on 

9 the national ambient air quality standards—unanimously recommended 

10 decreasing the primary standard to within the range of 0.060–0.070 ppm [60 to 

11 70 ppb]. It is the Committee’s consensus scientific opinion that your decision to 

12 set the primary ozone standard above this range fails to satisfy the explicit 

13 stipulations of the Clean Air Act that you ensure an adequate margin of safety for 

14 all individuals, including sensitive populations. 

15 In response to EPA’s solicitation of their advice on the Agency’s proposed rulemaking as 

16 part of the reconsideration, CASAC conveyed support (Samet, 2010). 

17 CASAC fully supports EPA’s proposed range of 0.060 – 0.070 parts per million 

18 (ppm) for the 8-hour primary ozone standard. CASAC considers this range to be 

19 justified by the scientific evidence as presented in the Air Quality Criteria for 

20 Ozone and Related Photochemical Oxidants (March 2006) and Review of the 

21 National Ambient Air Quality Standards for Ozone: Policy Assessment of 

22 Scientific and Technical Information, OAQPS Staff Paper (July 2007). As stated 

23 in our letters of October 24, 2006, March 26, 2007 and April 7, 2008 to former 

24 Administrator Stephen L. Johnson, CASAC unanimously recommended selection 

25 of an 8-hour average ozone NAAQS within the range proposed by EPA (0.060 to 

26 0.070 ppm). In proposing this range, EPA has recognized the large body of data 

27 and risk analyses demonstrating that retention of the current standard would 

28 leave large numbers of individuals at risk for respiratory effects and/or other 

29 significant health impacts including asthma exacerbations, emergency room 

30 visits, hospital admissions and mortality. 

31 In response to EPA’s request for additional advice on the reconsideration in 2011, 

32 CASAC reaffirmed their conclusion that “the evidence from controlled human and 

33 epidemiological studies strongly supports the selection of a new primary ozone standard within 

34 the 60 – 70 ppb range for an 8-hour averaging time” (Samet, 2011). As requested by EPA, 

35 CASAC’s advice and recommendations were based on the scientific and technical record from 

36 the 2008 rulemaking. In considering the record for the 2008 rulemaking, CASAC stated the 

37 following to summarize the basis for their conclusions (Samet, 2011, pp. ii to iii). 

38 The evidence available on dose-response for effects of ozone shows 

39 associations extending to levels within the range of concentrations 

40 currently experienced in the United States. 

41 There is scientific certainty that 6.6-hour exposures with exercise of 

42 young, healthy, non-smoking adult volunteers to concentrations ≥ 80 ppb 

43 cause clinically relevant decrements of lung function. 
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1 Some healthy individuals have been shown to have clinically relevant 

2 responses, even at 60 ppb. 

3 Since the majority of clinical studies involve young, healthy adult 

4 populations, less is known about health effects in such potentially ozone 

5 sensitive populations as the elderly, children and those with 

6 cardiopulmonary disease. For these susceptible groups, decrements in 

7 lung function may be greater than in healthy volunteers and are likely to 

8 have a greater clinical significance. 

9 Children and adults with asthma are at increased risk of acute 

10 exacerbations on or shortly after days when elevated ozone concentrations 

11 occur, even when exposures do not exceed the NAAQS concentration of 75 

12 ppb.

13 Large segments of the population fall into what EPA terms a “sensitive 

14 population group,’’ i.e., those at increased risk because they are more 

15 intrinsically susceptible (children, the elderly, and individuals with 

16 chronic lung disease) and those who are more vulnerable due to increased 

17 exposure because they work outside or live in areas that are more polluted 

18 than the mean levels in their communities. 

19 With respect to evidence from epidemiologic studies, CASAC stated “[W]hile epidemiological 

20 studies are inherently more uncertain as exposures and risk estimates decrease (due to the greater 

21 potential for biases to dominate small effect estimates), specific evidence in the literature does 

22 not suggest that our confidence on the specific attribution of the estimated effects of ozone on 

23 health outcomes differs over the proposed range of 60-70 ppb.” (Samet, 2011, p.10). 

24 In advice offered so far in the current review, which is considering an updated scientific 

25 and technical record since the 2008 rulemaking, CASAC has not yet conveyed their view on the 

26 adequacy of the current standard. In the first draft PA for the current review, staff reached the 

27 preliminary conclusion that the currently available evidence supports revising the standard to 

28 afford greater public health protection and that it does not support retention of the current 

29 standard (USEPA, 2012xx). Staff also concluded that the available evidence provides support 

30 for conducting further exposure and risk analyses of alternative standard levels in the range of 60 

31 to 70 ppb (USEPA, 2012xx). CASAC commented the draft PA provided “a strong scientific 

32 rationale for consideration of ozone levels (8 hour averages of 60 ppb to 70 ppb)” (Frey and 

33 Samet, 2012). 

34 3.4 PRELIMINARY STAFF CONCLUSIONS ON ADEQUACY OF PRIMARY 

35 STANDARD 

36 This section presents staff’s preliminary conclusions for the Administrator to consider in 

37 deciding whether it is appropriate to revise the existing primary O3 standard. Our conclusions are 

38 based on consideration of the assessment and integrative synthesis of the evidence presented in 
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1 the ISA, including consideration of air quality distributions in locations of selected 

2 epidemiologic studies; exposure and risk analyses in the second draft HREA; and the comments 

3 and advice of CASAC and public comment on earlier drafts of this document, and on the ISA 

4 and HREA. 

5 As an initial matter, staff concludes that reducing precursor emissions to achieve O3 

6 concentrations that meet the current standard will provide important improvements in public 

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

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

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

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

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

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

13 function decrements based on the information in these studies. Analyses in the second draft 

14 HREA estimate that the percentages of at-risk populations experiencing exposures of concern or 

15 abnormal and potentially adverse lung function decrements are substantially lower for air quality 

16 that just meets the current O3 standard than for recent air quality. 

17 Some support for this initial conclusion is also provided by estimates of O3-associated 

18 mortality and morbidity based on application of concentration-response relationships from 

19 epidemiologic studies to adjusted air quality. These estimates are more variable than estimates of 

20 O3 exposures and O3-induced lung function risks, and are associated with uncertainties that 

21 complicate their interpretation. However, epidemiology-based risk estimates for short- and long22 term O3 concentrations, in combination with the HREA’s national analysis of O3 responsiveness 

23 to reductions in NOX emissions and the larger body of health effects evidence, lead us to 

24 conclude that O3-associated mortality and morbidity would be expected to decrease following 

25 reductions in O3 precursor emissions to meet the current O3 standard. 

26 We next revisit the overarching policy question for this chapter, taking into consideration 

27 the responses to specific questions focused on the adequacy of the current primary O3 standard 

28 discussed above. 

29 Does the currently available scientific evidence and exposure/risk information, as 

30 reflected in the ISA and HREA, support or call into question the adequacy of the 

31 protection afforded by the current primary O3 standard? 

32 In considering the available evidence and information, staff concludes that the O3-

33 attributable health effects estimated to be allowed by air quality that meets the current primary 

34 standard for O3 can reasonably be judged important from a public health perspective. Thus, we 

35 conclude that the available health evidence and exposure/risk information call into question the 
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1 adequacy of the public health protection provided by the current standard. We further conclude 

2 that it is appropriate in this review to consider alternative standards that would increase public 

3 health protection, compared to the current standard, and that it is not appropriate to consider 

4 alternative standards with levels higher than the current standard, which would decrease public 

5 health protection (see chapter 4). The basis for these conclusions is discussed below. 

6 Studies evaluated since the completion of the 2006 O3 AQCD support and expand upon 

7 the strong body of evidence that, in the last review, indicated a causal relationship between short8 term O3 exposures and respiratory health effects. Together, experimental and epidemiologic 

9 studies support conclusions regarding a continuum of O3 respiratory effects ranging from small 

10 reversible changes in pulmonary function to more serious effects that can result in respiratory11 related emergency department visits, hospital admissions, and/or mortality. Recent animal 

12 toxicological studies support descriptions of modes of action for these respiratory effects and 

13 augment support for biological plausibility for the role of O3 in reported effects. With regard to 

14 mode of action, evidence indicates that antioxidant capacity may modify the risk of respiratory 

15 morbidity associated with O3 exposure. In addition, based on the consistency of findings across 

16 studies and evidence for the coherence of results from different scientific disciplines, strong 

17 evidence indicates that certain populations are at increased risk of O3-related effects. These 

18 include populations identified in previous reviews (i.e., people with asthma, children, older 

19 adults, outdoor workers) and populations identified since the last review (i.e., people with certain 

20 genotypes related to anti-oxidant and/or anti-inflammatory status; people with reduced intake of 

21 certain nutrients, such as Vitamins C and E). 

22 Evidence for adverse respiratory health effects attributable to “long-term” or repeated 

23 daily O3 exposures is much stronger than in previous reviews, and the ISA concludes there is 

24 likely to be a causal relationship between such O3 exposures and adverse respiratory health 

25 effects. Uncertainties related to the extrapolation of data generated by rodent toxicology studies 

26 to the understanding of health effects in humans have been reduced by studies in non-human 

27 primates and by recent epidemiologic studies. The evidence available in this review includes new 

28 epidemiologic studies using a variety of designs and analysis methods, conducted by different 

29 research groups in different locations, evaluating the relationships between long-term O3

30 exposures and measures of respiratory morbidity and mortality. New evidence supports 

31 associations between long-term or repeated O3 exposures and the development of asthma, with 

32 several studies reporting interactions between genetic variants and such O3 exposures. Studies 

33 also report associations between long-term or repeated O3 exposure and asthma prevalence, 

34 asthma severity and control, respiratory symptoms among asthmatics, and respiratory mortality. 

35 In considering the specific exposure concentrations reported to elicit respiratory effects, 

36 we note that recent evidence includes controlled human exposure studies reporting lung function 
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1 decrements and pulmonary inflammation in healthy adults engaged in intermittent, moderate 

2 exertion following 6.6 hour exposures to O3 concentrations as low as 60 ppb, and lung function 

3 decrements and respiratory symptoms following exposures to concentrations as low as 70 ppb. 

4 Compared to the evidence available in the last review, these studies have strengthened support 

5 for the occurrence of abnormal and potentially adverse respiratory effects following short-term 

exposures to O3 concentrations below 80 ppb.78 6 It is reasonable to judge exposures to such O3

7 concentrations to be potentially important from a public health perspective given the following: 

8 1. The respiratory effects reported following exposures to O3 concentrations of 60 and 70 

9 ppb, while at moderate exertion, can reasonably be judged adverse based on ATS criteria 

10 and past advice from CASAC. 

11

12 2. The controlled human exposure studies reporting these respiratory effects were conducted 

13 in healthy adults, while at-risk groups (e.g., asthmatics) could experience larger and/or 

14 more serious effects. 

15

16 3. These respiratory effects are coherent with the serious health outcomes that have been 

17 reported in epidemiologic studies (e.g., respiratory-related hospital admissions, 

18 emergency department visits, and mortality). 

19 Exposure estimates from the second draft HREA for urban case study areas indicate that, 

20 in areas just meeting the current O3 standard, approximately 10 to 20% of children would 

21 experience one or more exposures of concern to O3 concentrations of 60 ppb or above. In the 

22 case study areas evaluated in the HREA, this corresponds to over 2 million children experiencing 

23 approximately 4 million such exposures, including over 200,000 asthmatic children. Nationally, 

24 far more children would be expected to experience such exposures of concern in areas where the 

25 current standard is just met. On average over the years evaluated in the HREA, approximately 3 

26 to 8% of children are estimated to experience two or more exposures of concern to O3

27 concentrations of 60 ppb or greater. For the worst-case years (i.e., years with air quality patterns 

28 resulting in the highest exposure estimates), approximately 10 to 25% of children could 

29 experience one or more exposures of concern at or above 60 ppb, and up to 14% could 

30 experience two or more. 

31 Although the current standard more effectively limits exposures of concern at or above 

32 higher O3 concentrations (i.e., 70, 80 ppb), up to about 8% of children are estimated to 

33 experience exposures of concern at/above 70 ppb in the worst-case city and year (i.e., city and 

34 year with the largest estimates). In the worst-case city and year, about 2% of children were 

 

78 Cf. Misisssippi. 723 F. 3d at 262 (“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.”)
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1 estimated to experience two or more exposures of concern to O3 concentrations of 70 ppb or 

2 greater. 

3 Though we focus on children in these analyses of O3 exposures, we also recognize that 

4 exposures to 8-hour average O3 concentrations at or above 60, 70, or 80 ppb could be of concern 

5 for some adult populations. As discussed above, the patterns of exposure estimates over years 

6 and across cities are similar in adult asthmatics, older adults, and children, though smaller 

7 percentages of adult populations are estimated to experience exposures of concern. Thus, the 

8 results for children are one part of a broader range of potentially at-risk populations that also 

9 includes asthmatic adults and older adults. 

10 Consistent with estimates of exposures of concern, the second draft HREA also estimates 

11 that under air quality conditions just meeting the current O3 NAAQS, hundreds of thousands of 

12 asthmatic children would be expected to experience O3-induced lung function decrements that 

13 are large enough to be potentially adverse in people with lung disease. On average over the years 

14 evaluated in the HREA, the current standard is estimated to allow about 14% to 19% of children 

15 in urban case study areas, including asthmatic children, to experience one or more O3-induced 

16 lung function decrements > 10%. This corresponds to about 300,000 asthmatic children. 

17 Nationally, far more children would be expected to experience such O3-induced lung function 

18 decrements. About 8% to 12% of children are estimated to experience two or more decrements >

19 10%, on average. In the worst-case years, approximately 17% to 22% of children in the urban 

20 case study areas are estimated to experience one or more decrements > 10% and about 10% to 

21 14% are estimated to experience two or more such decrements. As with exposures of concern, 

22 the current standard more effectively limits larger O3-induced lung function decrements (i.e., >

23 15%, 20%). However, up to about 7% of children are estimated to experience one or more O3-

24 induced decrements > 15% in the worst-case city and year analyzed in the HREA (and as high as 

25 about 4% for two or more decrements). 

26 Recent epidemiologic studies also provide support, beyond that available in the last 

27 review, for associations between short-term O3 exposures and a wide range of adverse 

28 respiratory outcomes (including respiratory-related hospital admissions, emergency department 

29 visits, and mortality) and with total mortality. Associations with morbidity and mortality are 

30 stronger during the warm or summer months, and remain robust after adjustment for co31 pollutants. In one U.S. and several Canadian studies, associations with respiratory morbidity or 

32 mortality were reported in locations that would have met the current O3 standard. Even in some 

33 study locations where the current standard was not met, considering reported concentration34 response functions or cut-point analyses in the context of available air quality data indicate the 

35 existence of O3-health effect associations on the subsets of days with ambient O3 concentrations 

36 below 75 ppb. Taken together, these studies and associated air quality data indicate a relatively 
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1 high degree of confidence in the occurrence of O3-associated hospital admissions, emergency 

2 department visits, and mortality at ambient concentrations that meet the current standard. 

3 The HREA epidemiology-based risk estimates in 12 urban cases study areas indicate 

4 thousands of O3-associated hospital admissions, emergency department visits, and deaths per 

5 year for air quality conditions associated with meeting the current standard. Based on area-wide 

6 O3 concentrations from the upper portions of seasonal distributions, a focus that we judge 

7 appropriate given the greater certainty in O3-attributable effects at higher concentrations, 

8 hundreds to thousands of O3-associated deaths per year are estimated for air quality associated 

9 with the current standard in urban case study areas, indicating the potential for substantial public 

10 health risk. As recognized above in section 3.2.3.2, we note greater uncertainty in O3-attributable 

11 effects at lower concentrations, which are subject to increases upon air quality adjustment. 

12 Although there are additional uncertainties in quantifying risks by applying concentration13 response functions from epidemiology studies to adjusted O3 air quality, the general magnitude 

14 of risk estimates suggests the potential for a substantial number of O3-associated deaths and 

15 adverse respiratory events nationally when the current standard is met. 

16 In addition to the evidence and exposure/risk information discussed above, we also take 

17 note of the CASAC advice provided to the EPA Administrator on the proposed reconsideration 

18 of the 2008 decision establishing the current standard and the advice of the CASAC O3 Panel 

19 thus far in the current review. In commenting on the proposed reconsideration, the prior CASAC 

20 O3 Panel emphatically recommended revision of the standard to one with a lower level based 

21 entirely on the evidence and information in the record for the 2008 standard, which has been 

22 substantially strengthened in the current review (Samet, 2011; Samet, 2012). Based on review of 

23 the first draft PA in the current review, the current CASAC O3 Panel also described the draft PA 

24 as providing strong scientific rationale for consideration of lower standard levels (Frey and 

25 Samet, 2012). 

26 In consideration of all of the above, staff reaches the preliminary conclusion that the 

27 available evidence and exposure and risk information clearly calls into question the adequacy of 

28 public health protection provided by the current primary standard. This evidence and information 

29 provides strong support for the occurrence of a range of adverse respiratory effects, and 

30 mortality, under air quality conditions that would meet the current standard. Based on the 

31 analyses in the second draft HREA, we conclude that the exposures and risks projected to remain 

32 upon meeting the current standard are indicative of risks that can reasonably be judged to be 

33 important from a public health perspective. Thus, staff concludes that the evidence and 

34 information provides strong support for giving consideration to revising the current primary 

35 standard in order to provide increased public health protection against an array of adverse health 

36 effects that range from decreased lung function and respiratory symptoms to more serious 
3-123

1 indicators of morbidity (e.g., including emergency department visits and hospital admissions), 

2 and mortality. We further conclude that it is not appropriate to consider alternative standards 

3 with levels higher than the current standard, which would decrease public health protection. In 

4 consideration of all of the above, staff draws the preliminary conclusion that it is appropriate for 

5 the Administrator to consider revision of the current primary O3 standard to provide increased

6 public health protection. 

7

8
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1 3.5 REFERENCES 

2 Adams, W. C. (2006a) Comparison of chamber 6.6 hour exposures to 0.04-0.08 ppm ozone via square-wave and 

3 triangular profiles on pulmonary responses. Inhalation Toxicol. 18: 127-136. 

4 http://dx.doi.org/10.1080/08958370500306107 

5 Adams, WC. (2003a). Comparison of chamber and face mask 6.6-hour exposure to 0.08 ppm ozone via square-wave 

6 and triangular profiles on pulmonary responses. Inhal Toxicol 15: 265-281. 

7 Adams, WC. (2003b). Relation of pulmonary responses induced by 66-h exposures to 0.08 ppm ozone and 2-h 

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4-1

1 4 CONSIDERATION OF ALTERNATIVE PRIMARY STANDARDS 

2 Having reached the conclusion that the currently available scientific evidence and 

3 exposure/risk information calls into question the adequacy of the current O3 standard, we next 

4 consider the following overarching question: 

5 What is the range of potential alternative standards that are supported by the 

6 currently available scientific evidence and exposure/risk information, as reflected 

7 in the ISA and HREA respectively? 

8 To address this overarching question, in the sections below we evaluate a series of more specific 

9 questions related to the major elements of the NAAQS: indicator (section 4.1), averaging time 

10 (section 4.2), form (section 4.3), and level (section 4.4). In addressing these questions, we 

11 consider the currently available scientific evidence and exposure/risk information, including the 

12 evidence and information available at the time of the last review and that newly available in the 

13 current review, as assessed in the ISA and the second draft HREA. In so doing, we note that the 

14 final decision by the Administrator in this review will consider these elements collectively in 

evaluating the health protection afforded by the primary standard.115

16 4.1 INDICATOR 

17 In the last review, EPA focused on O3 as the most appropriate indicator for a standard 

18 meant to provide protection against ambient photochemical oxidants. In this review, while the 

19 complex atmospheric chemistry in which O3 plays a key role has been highlighted, no 

20 alternatives to O3 have been advanced as being a more appropriate indicator for ambient 

21 photochemical oxidants. More specifically, the ISA noted that O3 is the only photochemical 

22 oxidant (other than NO2) that is routinely monitored and for which a comprehensive database 

23 exists (ISA section 3.6). Data for other photochemical oxidants (e.g., PAN, H2O2, etc.) typically 

24 have been obtained only as part of special field studies. Consequently, no data on nationwide 

25 patterns of occurrence are available for these other oxidants; nor are extensive data available on 

26 the relationships of concentrations and patterns of these oxidants to those of O3 (ISA section 3.6). 

27 We further note that meeting an O3 standard can be expected to provide some degree of 

28 protection against potential health effects that may be independently associated with other 

29 photochemical oxidants, even though such effects are not discernible from currently available 

30 studies indexed by O3 alone. That is, since the precursor emissions that lead to the formation of 

31 O3 generally also lead to the formation of other photochemical oxidants, measures leading to 

 

1We 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 selecting a standard. Background O3 is discussed in more detail in chapter 2 of this 

second draft PA. 
4-2

1 reductions in population exposures to O3 can generally be expected to lead to reductions in 

2 population exposures to other photochemical oxidants. Taken together, we conclude that O3

3 remains the most appropriate indicator for a standard meant to provide protection against 

photochemical oxidants.24

5 4.2 AVERAGING TIME 

The EPA established the current 8-hour averaging time36 for the primary O3 NAAQS in 

7 1997 (62 FR 38856). The decision on averaging time in that review was based on numerous 

8 controlled human exposure and epidemiologic studies reporting associations between 6 to 8 hour 

9 O3 concentrations and adverse respiratory effects (62 FR 38861). It was also noted that a 

10 standard with a max 8-hour averaging time is likely to provide substantial protection against 

11 respiratory effects associated with 1-hour peak O3 concentrations. Similar conclusions were 

12 reached in the last O3 NAAQS review and thus, the 8-hour averaging time was retained in 2008. 

13 In the current review, we first consider the following question related to averaging time: 

14 To what extent does the available evidence continue to support the 

15 appropriateness of a standard with an 8-hour averaging time? 

16 In reaching conclusions related to this question, staff considers causality judgments from the 

17 ISA, as well as results from the specific controlled human exposure and epidemiologic studies 

18 that informed those judgments. These considerations are described below in more detail. 

19 As an initial consideration with respect to the most appropriate averaging time for the O3

20 NAAQS, we note that the strongest evidence for O3-associated health effects is for respiratory 

21 effects following short-term exposures. More specifically, the ISA concludes that evidence 

22 relating short-term O3 exposures to respiratory effects is “sufficient to infer a causal 

23 relationship.” The ISA also judges that short-term exposures to O3 are “likely to cause” both 

24 cardiovascular effects and mortality (U.S. EPA, 2013, section 2.5.2). Therefore, as in past 

25 reviews, the strength of the available scientific evidence provides strong support for a standard 

26 that protects the public health against short-term exposures to O3. 

27 In first considering the level of support available for specific short-term averaging times, 

28 we note the evidence available from controlled human exposure studies. As discussed in more 

29 detail in chapter 3 of this second draft PA, substantial health effects evidence from controlled 

30 human exposure studies demonstrates that a wide range of respiratory effects (e.g., pulmonary 

31 function decrements, increases in respiratory symptoms, lung inflammation, lung permeability, 

 

2The 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). 

3This 8-hour averaging time reflects daily max 8-hour average O3 concentrations. 
4-3

1 decreased lung host defense, and airway hyperresponsiveness) occur in healthy adults following 

2 6.6 hour exposures to O3 (EPA 2013, section 6.2.1.1). Compared to shorter exposure durations 

3 (e.g., 1-hour), studies evaluating 6.6 hour exposures in healthy adults have reported respiratory 

4 effects at lower O3 exposure concentrations and at more moderate levels of exertion.

5 We also note the strength of evidence from epidemiologic studies that have evaluated a 

6 wide variety of populations (e.g., including at-risk lifestages and populations, such as children 

7 and people with asthma, respectively). A number of different averaging times are used in O3

8 epidemiologic studies, with the most common being the max 1-hour concentration within a 24-

9 hour period (1-hour max), the max 8-hour average concentration within a 24-hour period (8-hr 

10 max), and the 24-hour average. These studies are discussed in chapter 3 of this second draft PA, 

11 and are assessed in detail in chapter 6 of the ISA (U.S. EPA, 2013). Limited evidence from time12 series and panel epidemiologic studies comparing risk estimates across averaging times does not 

13 indicate that one exposure metric is more consistently or strongly associated with respiratory 

14 health effects or mortality, though the ISA notes some evidence for “smaller O3 risk estimates 

15 when using a 24-hour average exposure metric” (EPA 2013, section 2.5.4.2; p. 2-31). For single16 and multi-day average O3 concentrations, lung function decrements were associated with 1-hour 

17 max, 8-hour max, and 24-hour average ambient O3 concentrations, with no strong difference in 

18 the consistency or magnitude of association among the averaging times (EPA 2013, p. 6-71). 

19 Similarly, in studies of short-term exposure to O3 and mortality, Smith et al. (2009) and Darrow 

20 et al. (2011) have reported high correlations between risk estimates calculated using 24-hour 

21 average, 8-hour max, and 1-hour max averaging times (EPA 2013, p. 6-253). Thus, the 

22 epidemiologic evidence alone does not provide a strong basis for distinguishing between the 

23 appropriateness of 1-hour, 8-hour, and 24-hour averaging times. 

24 Considering the health information discussed above, we conclude that an 8-hour 

25 averaging time remains appropriate for addressing health effects associated with short-term 

26 exposures to ambient O3. An 8-hour averaging time is similar to the exposure periods evaluated 

27 in controlled human exposure studies, including recent studies that provide evidence for 

28 respiratory effects following exposures to O3 concentrations below the level of the current 

29 standard. In addition, epidemiologic studies provide evidence for health effect associations with 

30 8-hour O3 concentrations, as well as with 1-hour and 24-hour concentrations. As in previous 

31 reviews, we note that a standard with an 8-hour averaging time (combined with an appropriate 

32 standard form and level) would also be expected to provide substantial protection against health 

33 effects attributable to 1-hour and 24-hour exposures (e.g., 62 FR 38861, July 18, 1997). 

34 The ISA also concludes that long-term O3 exposures are “likely to cause” respiratory 

35 effects (US EPA, 2013, chapter 7). Thus, in this review we also consider the extent to which 

36 currently available evidence and exposure/risk information suggests that a standard with an 8-
4-4

1 hour averaging time can provide protection against respiratory effects associated with longer

2 term exposures to ambient O3. In doing so, staff considers the following question: 

3 To what extent does the available evidence and exposure/risk information indicate 

4 that a standard with the current 8-hour averaging time could provide protection 

5 against long-term exposures to ambient O3?

6 In considering this issue in the last review of the O3 NAAQS, staff noted that “because long-term 

7 air quality patterns would be improved in areas coming into attainment with an 8-hr standard, the 

8 potential risk of health effects associated with long-term exposures would be reduced in any area 

9 meeting an 8-hr standard” (U.S. EPA, 2007, p. 6-57). 

10 In the current review, we further evaluate this issue, with a focus on the “long-term” O3

11 metrics reported to be associated with mortality or morbidity in recent epidemiologic studies. As 

12 discussed in section 3.1.3, much of the recent evidence for such associations is based on studies 

13 that defined long-term O3 in terms of seasonal averages of daily max concentrations (e.g., 

14 seasonal averages of 1-hour or 8-hour daily max concentrations). 

15 As an initial consideration, we note the risk results from the second draft HREA for 

16 respiratory mortality associated with long-term O3 concentrations. As discussed in section 

17 3.2.3.2, HREA analyses indicate that as air quality is adjusted to just meet the current 8-hour 

18 standard, most urban case study areas are estimated to experience reductions in respiratory 

19 mortality associated with long-term O3 concentrations based on the seasonal averages of 1-hour 

20 daily max O3 concentrations evaluated in the study by Jerrett et al. (2009) (U.S. EPA, 2014, 

21 chapter 7). As air quality is adjusted to meet lower potential alternative standard levels, for 

22 standards based on 3-year averages of the annual fourth-highest daily max 8-hour O3

23 concentrations, respiratory mortality risks are estimated to be reduced further in urban case study 

24 areas (section 4.4.2.3,below). This analysis indicates that an O3 standard with an 8-hour 

25 averaging time, when coupled with an appropriate form and level, can reduce respiratory 

26 mortality reported to be associated with “long-term” O3 concentrations. 

27 In further considering the study by Jerrett et al. (2009), we compare long-term O3

28 concentrations following model adjustment in urban case study areas (i.e., adjusted to meet the 

29 current and potential alternative 8-hour standards) to the concentrations present in study cities 

30 that provided the basis for the positive and statistically significant association with respiratory 

31 mortality. As indicated below (Table 4-3), this comparison suggests that a standard with an 8-

32 hour averaging time can decrease seasonal averages of 1-hour daily max O3 concentrations, and

33 can maintain those O3 concentrations below the seasonal average where we have the most 

34 confidence in the reported concentration-response relationship with respiratory mortality (see 

35 section 4.4.1 for further discussion). 
4-5

1 The second draft HREA also conducted analyses evaluating the impacts of reducing 

regional NOX emissions on the seasonal averages of 8-hour daily max O3 concentrations.42

3 Seasonal averages of 8-hour daily max O3 concentrations reflect long-term metrics that have 

4 been reported to be associated with respiratory morbidity effects in several recent O3

5 epidemiologic studies (e.g., Islam et al., 2008; Lin et al., 2008; Salam et al., 2009). The HREA 

6 analyses indicate that the large majority of the U.S. population lives in locations where reducing 

7 NOX emissions would be expected to result in decreases in seasonal averages of daily max 8-

8 hour ambient O3 concentrations (U.S. EPA, 2014, chapter 8). Thus, consistent with the 

9 respiratory mortality risk estimates noted above, this analysis suggests that reductions in O3

10 precursor emissions in order to meet a standard with an 8-hour averaging time would also be 

11 expected to reduce the types of long-term O3 concentrations that have been reported in recent 

12 epidemiologic studies to be associated with respiratory morbidity. 

13 Taken together, we conclude that a standard with an 8-hour averaging time, coupled with 

the current 4th 14 high form and an appropriate level, would be expected to provide appropriate 

15 protection against the long-term O3 concentrations that have been reported to be associated with 

16 respiratory morbidity and mortality. This issue is considered further, within the context of 

17 specific potential alternative standard levels, in section 4.4 below. 

18 4.3 FORM 

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

20 level of the standard in determining whether an area attains the standard. The foremost 

21 consideration in selecting a form for potential alternative primary standards is the adequacy of 

22 the public health protection provided by the combination of the form and the other elements of 

23 the standard. As such, in reaching staff conclusions regarding the appropriate form(s) to consider 

24 for a potential alternative primary O3 standard, we consider the following question: 

25 To what extent do the available evidence and/or information continue to support 

26 the appropriateness of a standard with a form defined by the 3-year average of 

annual 4th 27 -highest 8-hour daily max O3 concentrations?

28 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.529 An advantage of the 

30 current concentration-based form recognized in the 1997 review is that such a form better 

31 reflects the continuum of health effects associated with increasing ambient O3 concentrations. 

 

4Analyses 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. 

5

For a standard with a 1-expected-exceedance form to be met at an air quality monitoring site, the fourthhighest air quality value in 3 years, given adjustments for missing data, must be less than or equal to the level of the 

standard. 
4-6

1 Unlike an expected exceedance form, a concentration-based form gives proportionally more 

2 weight to years when 8-hour O3 concentrations are well above the level of the standard than to 

3 years when 8-hour O3 concentrations are just above the level of the standard. It was judged 

4 appropriate to give more weight to higher O3 concentrations, given that available health evidence 

5 indicated a continuum of effects associated with exposures to varying concentrations of O3, and 

6 given that the extent to which public health is affected by exposure to ambient O3 is related to the 

7 actual magnitude of the O3 concentration, not just whether the concentration is above a specified 

8 level. 

9 During the 1997 review, EPA considered a range of alternative “concentration-based” 

10 forms, including the second-, third-, fourth- and fifth-highest daily max 8-hour concentrations in 

11 an O3 season. The fourth-highest daily max was selected, recognizing that a less restrictive form 

12 (e.g., fifth highest) would allow a relatively large percentage of sites to experience O3 peaks well 

13 above the level of the standard, and would allow more days on which the level of the standard 

14 may be exceeded when attaining the standard (62 FR 38856). Consideration was also given to 

15 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.616 A 

17 more restrictive form was not selected, recognizing that the differences in the degree of 

18 protection afforded by the alternatives were not well enough understood to use any such 

19 differences as a basis for choosing the most restrictive forms (62 FR 38856). 

20 In the 2008 review, EPA additionally considered the potential value of a percentile-based 

21 form. In doing so, EPA recognized that such a statistic is useful for comparing datasets of 

22 varying length because it samples approximately the same place in the distribution of air quality 

23 values, whether the dataset is several months or several years long. However, EPA concluded 

24 that a percentile-based statistic would not be effective in ensuring the same degree of public 

25 health protection across the country. Specifically, a percentile-based form would allow more 

26 days with higher air quality values in locations with longer O3 seasons relative to places with 

27 shorter O3 seasons. 

28 Thus, in the 2008 review EPA concluded that a form based on the nth-highest max O3

29 concentration would more effectively ensure that people who live in areas with different length 

30 O3 seasons receive the same degree of public health protection. Based on analyses for forms 

31 specified in terms of an nth-highest concentration (n ranged from 3 to 5), advice from CASAC,

 

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

and public comment,71 the Administrator concluded that a 4th-highest daily max should be 

2 retained (73 FR 16465). In reaching this decision, the Administrator recognized that “the 

3 adequacy of the public health protection provided by the combination of the level and form is a 

4 foremost consideration” (73 FR 16475). 

5 The Administrator also recognized that it is important to have a form that provides 

6 stability with regard to implementation of the standard. In the case of O3, for example, he noted 

7 the importance of a form insulated from the impacts of the meteorological events that are 

8 conducive to O3 formation. Such events could have the effect of reducing public health 

9 protection, to the extent they result in frequent shifts between meeting and violating the standard 

10 due to meteorological conditions. The Administrator noted that such frequent shifting could 

11 disrupt an area’s ongoing implementation plans and associated control programs (73 FR 16474).

12 In its notice of proposed rulemaking to reconsider the 2008 standard, the EPA did not propose to 

13 reconsider the form of the standard. 

14 In the current review, we consider the extent to which newly available information 

15 provides support for consideration of alternative forms. In so doing, we take note of the 

16 conclusions of prior reviews summarized above. We recognize the value of an nth-high statistic 

17 over that of an expected exceedance or percentile-based form in the case of the O3 standard, for 

18 the reasons summarized above. We additionally take note of the importance of stability in 

19 implementation to achieving the level of protection specified by the NAAQS. Specifically, we 

20 note that to the extent that areas engaged in implementing the O3 NAAQS frequently shift from 

21 meeting to violating the standard, it is possible that ongoing implementation plans and associated 

22 control programs could be disrupted, thereby reducing public health protection.

23 In light of this, while giving foremost consideration to the adequacy of public health 

24 protection provided by the combination of all elements of the standard, including the form, we 

25 consider particularly findings from prior reviews with regard to the use of the nth-high metric. 

As noted above, the 4th 26 -highest daily max was selected in 1997 in recognition of the public 

27 health protection provided by this form, when coupled with an appropriate averaging time and 

28 level, and recognizing that such a form can provide stability for implementation programs. The 

29 currently available evidence and information does not call into question these conclusions from 

previous reviews. Therefore, we conclude that it would be appropriate to retain the current 4th 30 -

 

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

1 highest daily max form for an O3 standard with an 8-hour averaging time and a revised level, as 

2 discussed below. 

3 4.4 LEVEL 

4 In considering potential alternative standards levels to provide greater protection than that 

5 afforded by the current standard against O3-related adverse health effects, we address the 

6 following overarching question. 

7 For an O3 standard defined in terms of the current indicator, averaging time, and 

8 form, what alternative levels are appropriate to consider in order to provide 

9 adequate public health protection against short- and long- term exposures to O3

10 in ambient air? 

11 In considering this question, we take into account the experimental and epidemiologic evidence 

12 as presented in the ISA, as well as the uncertainties and limitations associated with this evidence 

13 (section 4.4.1). In addition, we consider the quantitative estimates of exposure and risk provided 

14 by the HREA, as well as the uncertainties and limitations associated with these risk estimates 

15 (section 4.4.2). 

16 4.4.1 Evidence-based Considerations 

17 In this section, we consider the available evidence from controlled human exposure and 

18 epidemiologic studies, including the uncertainties and limitations associated with that evidence, 

19 within the context of potential alternative standard levels. We consider both the exposure 

20 concentrations at which controlled human exposure studies provide evidence for health effects,

21 and the ambient O3 concentrations present in locations where epidemiologic studies have 

22 reported health effect associations (see also section 3.1). 

23 Controlled human exposure studies 

24 We consider the following question related to controlled human exposure studies: 

25 To what extent does the available evidence from controlled human exposure 

26 studies provide support for consideration of potential alternative standard levels 

27 lower than 75 ppb? 

28 To inform our conclusions regarding this question, we consider the lowest O3 concentrations at 

29 which various effects have been evaluated and statistically significant effects reported. We also 

30 consider the potential for reported effects to be adverse, including in at-risk populations. 

31 As discussed in section 3.1.2.1, in healthy adults group mean O3-induced lung function 

32 decrements exhibit a smooth dose-response relationship without evidence of a threshold from 40 

33 to 120 ppb O3 (US EPA, 2013, Figure 6-1). The lowest O3 exposure concentration for which 

34 statistically significant decrements have been reported is 60 ppb (Brown, 2006; Kim et al., 2011). 
4-9

1 The ISA concludes that mean FEV1 is clearly decreased by 6.6-hour exposures to O3

2 concentrations of 60 ppb and higher in young, healthy adults during moderate exertion (US EPA, 

3 2013, p. 6-9). As discussed in section 3.1.3, such a decrease in mean lung function meets the 

4 ATS criteria for an adverse response given that a downward shift in the distribution of FEV1

5 would result in diminished reserve function, and therefore would increase risk from further 

6 environmental insult. In addition, the ISA notes that following exposures to 60 ppb O3 10% of 

7 healthy individuals experience FEV1 decrements > 10% (U.S. EPA, 2013, page 6-19). A 10% 

8 decrement in FEV1 is accepted by ATS as an abnormal response, and based on advice received 

9 from CASAC in previous reviews, such decrements could be adverse in people with lung disease 

10 (section 3.1.3). 

11 As discussed in section 3.1.2.1, one recent controlled human exposure study has reported 

12 O3-induced pulmonary inflammation (PMN influx to the ELF) following exposures of young, 

13 healthy adults to O3 concentrations of 60 ppb (Kim et al., 2011), the lowest concentration at 

14 which inflammatory responses have been evaluated in human studies. Induction of pulmonary 

15 inflammation is evidence that injury has occurred. The possibility of chronic effects due to 

16 repeated inflammatory events has been evaluated in animal studies. Repeated events of acute 

17 inflammation can have several potentially adverse outcomes including: induction of a chronic 

18 inflammatory state; altered pulmonary structure and function, leading to diseases such as asthma; 

19 altered lung host defense response to inhaled microorganisms, particularly in potentially at-risk 

20 populations such as the very young and old; and, altered lung response to other agents such as 

21 allergens or toxins (U.S. EPA, 2013, Section 6.2.3). Thus, lung injury and the resulting 

22 inflammation, particularly if experienced repeatedly, provide a mechanism by which O3 may 

23 cause other more serious respiratory effects (e.g., asthma exacerbations) and possibly 

24 extrapulmonary effects. 

25 With respect to respiratory symptoms, a recent study by Schelegle et al. (2009) reported a 

26 statistically significant increase in respiratory symptoms in young, healthy adults following 6.6 

27 hour exposures to an average O3 concentration of 70 ppb. This study also reported a statistically 

28 significant decrease in FEV1 following such exposures. As discussed in section 3.1.3, the 

29 occurrence of both lung function decrements and respiratory symptoms meets criteria established 

30 by the ATS defining an “adverse” respiratory response. Although some studies have reported 

31 that respiratory symptoms develop during exposures at 60 ppb, the increases in symptoms in 

32 these studies have not reached statistical significance by the end of the 6.6 hr exposures (Adams 

2006; Schelegle et al., 2009).833

 

8Adams (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. 
4-10

1 Based on the results discussed above and in section 3.1.2.1, we conclude that controlled 

2 human exposure studies provide evidence of potentially adverse lung function decrements and 

3 airway inflammation in healthy individuals following exposures to 60 ppb O3, and evidence of 

4 respiratory symptoms combined with lung function decrements (an “adverse” response based on 

5 ATS criteria) following exposures to 70 ppb. In reaching these conclusions, we recognize that 

6 most studies have not evaluated exposure concentrations below 60 ppb, and that 60 ppb does not 

7 necessarily reflect an exposure concentration below which effects no longer occur. Specifically, 

8 given the occurrence of airway inflammation following exposures to 60 ppb and higher, it may 

9 be reasonable to expect that inflammation would also occur following exposures to O3

10 concentrations somewhat below 60 ppb. Although some studies show that respiratory symptoms 

11 develop during exposures at 60 ppb, they have not reached statistical significance by the end of 

12 the 6.6 hr exposures (Adams 2006; Schelegle et al. 2009). Thus, respiratory symptoms combined 

13 with lung function decrements are likely to occur to some degree in healthy individuals with 6.6-

14 hr exposures to concentrations below 70 ppb. Further, we note that these controlled human 

15 exposure studies were conducted in healthy adults and that people with asthma, including 

16 asthmatic children, are likely to be more sensitive to O3-induced respiratory effects. Therefore, 

17 these exposure concentrations are more likely to cause adverse respiratory effects in children and 

18 adults with asthma, and more generally in people with respiratory disease. 

19 In further considering effects following exposures to O3 concentrations below 75 ppb, in 

20 section 3.1.4.1 we discuss panel studies highlighted in the ISA for the extent to which monitored 

21 ambient O3 concentrations reflect exposure concentrations in their study populations (U.S. EPA, 

22 2013, section 6.2.1.2). These panel studies used on-site monitoring to evaluate O3-attributable 

23 lung function decrements in individuals engaged in outdoor recreation, exercise, or work. Table 

24 3-2 includes O3 panel studies that report analyses of O3-attributable lung function decrements for 

25 O3 concentrations at or below 75 ppb, and that measure O3 concentrations with monitors located 

26 in the areas where study subjects were active (e.g., on site at summer camps or in locations 

27 where exercise took place). Consistent with the results of controlled human exposure studies 

28 discussed above, these panel studies report associations with lung function decrements for 

29 subjects exposed to on-site monitored O3 concentrations below 75 ppb. Associations in panel 

30 studies have been reported for a wider range of populations than has been evaluated in controlled 

31 human exposure studies, including children. 

32 In considering controlled human exposure studies of other O3-induced effects, we note 

33 that airway hyper-responsiveness and impaired lung host defense capabilities have been reported 

34 in healthy adults engaged in moderate exertion following exposures to O3 concentrations as low 

35 as 80 ppb, the lowest concentration evaluated for these effects. As discussed in section 3.1.2.1, 

36 these physiological effects have been linked to aggravation of asthma and increased 
4-11

1 susceptibility to respiratory infection, potentially leading to increased medication use, increased 

2 school and work absences, increased visits to doctors’ offices and emergency departments, and 

3 increased hospital admissions. These are all indicators of adverse O3-related morbidity effects, 

4 which are consistent with, and provide plausibility for, the adverse morbidity effects and 

5 mortality effects observed in epidemiologic studies. 

6 In revisiting the question above, we conclude that the available controlled human 

7 exposure evidence supports an upper end of the range of potential alternative standard levels for 

8 consideration no higher than 70 ppb. For 6.6 hour exposures at 70 ppb, lung function decrements 

9 and respiratory symptoms, a combination of effects that meet ATS criteria for an adverse 

10 response (as discussed in section 3.1.3), have been demonstrated in healthy adults in controlled 

human exposure studies.9

11 In addition, potentially adverse respiratory effects, including lung 

12 function decrements and airway inflammation, have been demonstrated following 6.6 hour 

13 exposures to O3 concentrations below 70 ppb (i.e., at 60 ppb, as discussed below). A level of 70 

14 ppb would also be below the lowest-observed-effects level for effects such as airway hyper15 responsiveness and impaired host-defense capabilities in healthy adults while at prolonged 

16 moderate exertion. As discussed in section 3.1.2.1 of this second draft PA, such physiological 

17 effects have been linked to aggravation of asthma and increased susceptibility to respiratory 

18 infection, potentially leading to increased medication use, increased school and work absences, 

19 increased visits to doctors’ offices and emergency departments, and increased hospital 

20 admissions. 

21 Based on the above considerations, we also conclude that the evidence from controlled 

22 human exposure studies supports setting the lower end to the range of alternative O3 standards at 

23 60 ppb. Potentially adverse lung function decrements and pulmonary inflammation have been 

24 demonstrated to occur at in healthy adults at 60 ppb. This is a short-term exposure concentration 

25 that may be reasonably concluded to elicit adverse effects in at-risk groups. Pulmonary 

26 inflammation, particularly if experienced repeatedly, provides a mechanism by which O3 may 

27 cause other more serious respiratory morbidity effects (e.g., asthma exacerbations) and possibly 

28 extrapulmonary effects. 

29 Epidemiologic evidence 

30 We also consider what the information from epidemiologic studies indicates with regard 

31 to potential alternative standard levels appropriate for consideration. Based on the information in 

32 section 3.1.4.2 of this second draft PA (see Table 3-3), we first note that several epidemiologic 

 

9Although some studies report that respiratory symptoms develop during exposures to 60 ppb O3, these effects have 

not reached statistical significance by the end of the 6.6 hour exposures (Adams, 2006; Schelegle et al., 2009). Thus 

respiratory symptoms, in combination with lung function decrements, are likely to occur to some degree in healthy 

individuals following exposures to O3 concentrations somewhat below 70 ppb. 
4-12

1 studies have reported positive and statistically significant associations with hospital admissions, 

2 emergency department visits, and/or mortality in study areas where ambient O3 concentrations 

3 would have met the current standard (i.e., with its level of 75 ppb). This includes Canadian 

4 multicity studies in which the majority of study cities would have met the current standard over 

5 entire study periods (Cakmak et al., 2006; Dales et al., 2006; Katsouyanni et al., 2009; Stieb et 

6 al., 2009), and a U.S. single-city study conducted in a location likely to have met the current 

7 standard over the entire study period (Mar and Koenig, 2009). 

8 In further evaluating these studies, and building upon our conclusions based on controlled 

9 human exposures studies, as discussed above, we consider the following question related to the 

10 epidemiologic evidence: 

11 To what extent have U.S. and Canadian epidemiologic studies reported associations 

12 with mortality or morbidity in locations likely to have met potential alternative O3

13 standards with levels from 70 to 60 ppb?

14 Addressing this question can provide important insights into the extent to which O3-

15 health effect associations are present for distributions of ambient O3 concentrations that would be 

16 allowed by various potential alternative standards. To the extent O3 health effect associations are 

17 reported in study areas that would have met potential alternative standards, we have greater 

18 confidence that exposures to ambient O3 concentrations allowed by such alternatives could result 

in the types of clearly adverse effects evaluated in these studies.10 19 Therefore, our focus in this 

20 section is to consider what these studies convey regarding the extent to which health effects may 

21 be occurring (i.e., as indicated by associations) under air quality conditions allowed by potential 

22 alternative standards. Specifically, we consider the numbers of study locations likely to have met 

23 potential alternative standards with levels of 70, 65, and 60 ppb during study periods (Table 4-1). 

24

 

10See ATA III, 283 F.3d at 370 (EPA justified in revising NAAQS when health effect associations are observed at 

levels allowed by the NAAQS). 
4-13

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

2 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

3 No U.S. or Canadian studies reported positive and statistically significant health effect 

4 associations when all study locations would have met a standard with a level from 70 to 60 ppb 

5 over the entire study period. However, for the studies by Cakmak et al. (2006), Katsouyanni et 

6 al. (2009), and Stieb et al. (2009), the majority of study locations would likely have met a 

7 standard with a level of either 70 or 65 ppb (Cakmak et al., 2006; Katsouyanni et al., 2009; Stieb 

8 et al., 2009). In contrast, the majority of locations in these studies would likely have violated a 

9 standard with a level of 60 ppb. While there is uncertainty in ascribing the multicity effect 

10 estimates reported in these Canadian studies to ambient concentrations that would have met 

11 standards with levels of 70 or 65 ppb (i.e., given that some study locations would have violated 

12 such standards over at least part of the study period), reported multicity effect estimates are 

13 largely influenced by locations meeting these potential alternative standards. 

14 As with our consideration of the current standard (section 3.1.4.2), we next consider the 

15 extent to which epidemiologic studies have characterized O3 health effect associations, including 

16 confidence in those associations, for various portions of distributions of ambient O3

17 concentrations. In considering such analyses within the context of potential alternative standards, 
4-14

1 we focus on the extent to which epidemiologic studies report health effect associations for air 

2 quality distributions restricted to ambient pollutant concentrations below one or more 

3 predetermined cut-points. As discussed in section 3.1.4.2, such “cut-point” analyses can provide 

4 information on the magnitude and statistical precision of effect estimates for defined 

5 distributions of ambient concentrations, which may in some cases include distributions that 

6 would be allowed by potential alternative standards. Specifically, we consider the following 

7 question: 

8 To what extent do cut-point analyses from epidemiologic studies report health effect 

9 associations at ambient O3 concentrations that are likely to be allowed by potential 

10 alternative standards with levels from 70 to 60 ppb?

11 As with our consideration of the current standard in section 3.1.4.2 of this second draft 

12 PA, we evaluate the cut-point analyses presented in the U.S. multicity study by Bell et al. (2006). 

13 These cut-point analyses can provide insights into the magnitude and statistical precision of 

14 health effect associations for different portions of the distribution of ambient concentrations, 

15 including insights into the ambient concentrations below which uncertainty in reported 

16 associations becomes notably greater. Our analysis of air quality data associated with the cut17 points evaluated by Bell et al., and uncertainties associated with that analysis, is described 

18 elsewhere in this document (section 3.1.4.2). In this section, we consider what these cut-point 

19 analyses indicate with regard to the potential for health effect associations to extend to ambient 

20 O3 concentrations likely to be allowed by a revised O3 NAAQS with a level below 75 ppb. 

21 We particularly focus on the lowest cut-point for which the association between O3 and

22 mortality was reported to be statistically significant (i.e., 30 ppb, as discussed in section 3.1.4.2). 

23 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 24 highest 8-hour daily max O3

25 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 26 highest 8-hour daily max O3 concentrations at or 

below 65 ppb (Table 4-2). In contrast, the majority of study areas had 4th 27 highest concentrations 

28 above 60 ppb. While there are uncertainties in interpreting these cut-point analyses within the 

29 context of potential alternative standard levels, they suggest that the majority of the air quality 

30 distributions that provided the basis for a positive and statistically significant association with 

31 mortality would have been allowed by a standard with a level of 70 or 65 ppb, but would have 

32 violated a standard with a level of 60 ppb. For higher cut-points, all of which also resulted in 

33 statistically significant associations with mortality, the majority of study cities had 3-year 

averages of annual 4th 34 highest 8-hour daily max concentrations greater than 70 ppb. 

35
4-15

Table 4-2. Number of study cities with 3-year averages of 4th 1 highest 8-hour daily max

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

3 presented in Bell et al. (2006) 

Cut-point for 2-day moving average across monitors and cities (24-h avg)

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%)

4 In considering the implications of these analyses for potential alternative standard levels, 

5 we also note the important uncertainties described in section 3.1.4. Several of these uncertainties 

6 become increasingly important as health effect associations are evaluated for lower ambient O3

7 concentrations, such as when considering associations reported at the lower ends of the 

8 distributions of ambient O3. These include uncertainties that could obscure presence of potential 

9 thresholds, affecting our characterization of confidence in O3 health effect associations over 

10 distributions of ambient concentrations; uncertainty in the extent to which the relatively low 

11 ambient O3 concentrations present in some study areas cause or contribute to reported effects; 

12 and uncertainty in the extent to which we were able to identify the air quality data associated 

13 with health effects in some published analyses (particularly for the subset analyses by Bell et al., 

14 2006) (section 3.1.4.2). 

15 We next consider the extent to which epidemiologic studies employing longer-term 

16 ambient O3 concentration metrics can inform our consideration of potential alternative standard 

17 levels. In doing so, we consider the following question: 

18 To what extent does the available evidence indicate that a standard with a level 

from 70 to 60 ppb, combined with the current 8-hour averaging time and 4th 19 high 

20 form, could provide protection from long-term exposures to ambient O3

21 concentrations for which there is evidence of health effects? 

22 We first note that, as discussed in section 3.1.4.3 of this second draft PA, virtually all of 

23 the study cities that provided the basis for the positive and statistically significant association 
4-16

1 between long-term O3 and respiratory mortality (Jerrett et al., 2009) would have violated the 

2 current standard, and therefore potential alternative standards with lower levels. Thus, as with 

3 our consideration of the current standard in section 3.1.4.3, while the study by Jerrett et al. 

4 (2009) contributes to our understanding of health effects associated with ambient O3

5 (summarized in section 3.1.2), it is less informative regarding the extent to which those health 

6 effects may be occurring under air quality conditions that would meet potential alternative 

7 standards. 

8 To further evaluate this issue, we use the adjusted air quality in urban case study areas, as 

9 described in the second draft HREA, to consider the extent to which just meeting alternative O3

10 standards with levels of 70, 65, and 60 ppb could maintain long-term O3 concentrations below 

11 those in the cities that provided the basis for the positive and statistically significant association 

with respiratory mortality reported by Jerrett et al. (2009).11 12 Upon adjustment of air quality in 

13 U.S. urban case study areas to meet the current and potential alternative 8-hour standards, 

14 seasonal average 1-hour daily max concentrations were calculated and compared to the 

15 concentrations in study cities.

16 As discussed in section 3.1.4.3, Jerrett et al. (2009) reported that when seasonal averages 

of 1-hour daily max O3 concentrations12 17 ranged from 33 to 104 ppb, there was no statistical 

18 deviation from a linear concentration-response relationship between O3 and respiratory mortality 

19 across 96 U.S. cities (U.S. EPA, 2013, section 7.7). However, as discussed in section 3.1.4.3, we 

20 have the greatest confidence in the reported linear concentration-response function for “long21 term” O3 concentrations above the first quartile (i.e., 53.2 ppb), given the notable widening in 

22 confidence intervals for lower concentrations (based on visual inspection of Figure 3-6 in section 

3.1.4.3); the limited evidence noted by study authors for a threshold at 56 ppb;13 23 and the fact that 

24 most study cities contributing to the linear function had O3 concentrations in the highest three 

25 quartiles (accounting for approximately 72% of the respiratory deaths in the cohort, based on 

26 Table 2 in the published study). 

27 Given the above, we note the extent to which long-term O3 concentrations (i.e., seasonal 

28 average of 1-hour daily max) in urban case study areas are estimated to be at or below 53 ppb 

 

11Air 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 second draft 

HREA (chapter 4). After a given adjustment, seasonal average 1-hour daily max concentrations were calculated. 

12Jerrett 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. 

13The ISA does not reach conclusions regarding the potential for a threshold in the association between “long-term” 

O3 concentrations and respiratory mortality. 
4-17

1 following model adjustment to meet potential alternative standards with levels of 70, 65, and 60 

2 ppb. To the extent air quality adjustment to just meet potential alternative short-term standards 

3 results in long-term concentrations near or below 53 ppb, we have greater confidence in the 

4 degree to which those short-term standards could protect against the health effects associated 

5 with longer term O3 exposures. Though there is uncertainty associated with these comparisons 

6 (e.g., due to uncertainty in the potential for a threshold to exist; uncertainty in the long-term 

7 concentration below which confidence intervals widen notably, based on visual inspection of 

8 concentration-response function in the published study; and the limited number of urban case 

9 study areas for which adjusted air quality is available), this analysis can provide insight into the 

10 extent to which various alternative short-term standards would be expected to maintain long-term 

11 O3 concentrations below those where we have the most confidence in the reported concentration12 response relationship with respiratory mortality. 

13 Table 4-3 indicates that when considering recent (i.e., unadjusted) air quality, 2 of 12 

14 urban case study areas had seasonal average 1-hour daily max O3 concentrations at or below 53

15 ppb in all of the years examined. When air quality was adjusted to just meet the current 8-hour 

16 standard (75 ppb in Table 4-3), 6 of 12 urban case study areas had seasonal average 1-hour daily 

17 max O3 concentrations at or below 53 ppb in all of the years examined. When air quality is 

18 further adjusted to just meet potential alternative standards with lower levels, seasonal averages 

19 of 1-hour daily max O3 concentrations are estimated to be at or below 53 ppb in 9 of 12 urban 

20 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).14 21 Though as noted above there are important uncertainties 

22 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 23 high form 

24 and an appropriate standard level, would be expected to maintain seasonal averages of 1-hour 

25 daily max O3 concentrations below those where analyses indicate the most confidence in the 

26 concentration-response relationship with respiratory mortality reported by Jerrett et al. (2009). 

 

14As described in the second draft HREA, a standard level of 60 ppb was not evaluated in New York City (U.S. 

EPA, 2014, chapter 4). 
4-18

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

2 study areas for recent air quality and air quality adjusted to just meet the 

3 current and potential alternative standards. 

4

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 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 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
4-19

1 Based on the above analyses, we conclude that the available epidemiologic evidence is 

2 consistent with the available evidence from controlled human exposure studies in providing 

3 support for consideration of a standard level in the range of 70 to 60 ppb. Compared to the 

4 current standard, a standard level from within this range would expected to be more effective at 

5 maintaining short-term and long-term ambient O3 concentrations below those where the evidence 

6 indicates O3-associated mortality and/or morbidity. 

7 In reaching overall staff conclusions about an appropriate range of standard levels for 

8 consideration, we further evaluate the results of the exposure and risk assessments that are based 

9 on modeling changes in the entire distribution of ambient O3 concentrations to simulate just 

10 meeting potential alternative standards. These results are discussed below in section 4.4.2. 

11 4.4.2 Air Quality-, Exposure-, and Risk-Based Considerations 

12 Beyond considering the available evidence, we also consider the extent to which specific 

13 potential alternative standard levels, in conjunction with the current averaging time and form (3-

year average of annual 4th 14 highest 8-hour daily max), could reduce estimated O3 exposures and 

15 health risks. In the first draft PA (U.S. EPA, 2012b), we concluded that the available evidence 

16 supports conducting further exposure and risk analyses of potential alternative standard levels in 

17 the range of 70 down to 60 ppb. Based on these conclusions, the second draft HREA evaluates 

18 exposures and risks estimated to be associated with potential alternative standard levels from the 

19 upper (70 ppb), middle (65 ppb), and lower (60 ppb) portions of this range. In considering these 

20 analyses in this second draft PA, we consider the following question: 

21 To what extent does the available exposure and risk information provide support 

22 for considering potential alternative standard levels from 70 to 60 ppb, when 

combined with the current 8-hour averaging time and 4th 23 high form? 

24 In considering exposure and risk analyses, we emphasize the nature and magnitude of the O3

25 exposures and health risks estimated to remain upon just meeting each alternative standard level, 

26 and the changes in exposures and risks estimated for each alternative level when compared to the 

27 current standard. Section 4.4.2.1 below discusses our exposure-based considerations. Sections 

28 4.4.2.2 and 4.4.2.3 discuss our consideration of estimates of lung function risks and estimates of 

29 epidemiology-based mortality/morbidity risks, respectively. 

30 4.4.2.1 Exposure-Based Considerations 

31 As discussed in more detail in section 3.2.2 of this second draft PA, the exposure 

32 assessment presented in the second draft HREA (U.S. EPA, 2014) provides estimates of the 

33 number and percent of people exposed to O3 concentrations at or above benchmark 

34 concentrations of 60, 70, and 80 ppb, while at moderate or greater exertion. Estimates of such 
4-20

1 “exposures of concern” provide perspective on the potential public health impacts of O3-related 

2 effects, including for effects that cannot currently be evaluated in a quantitative risk assessment. 

3 The approach taken in the second draft HREA to estimating exposures of concern, and the key 

4 uncertainties associated with exposure estimates, are summarized in section 3.2.2 for air quality 

5 adjusted to just meet the current standard and are discussed in more detail in chapter 5 of the 

6 second draft HREA (U.S. EPA, 2014). As discussed in section 3.2.2, when evaluating potential 

7 alternative standard levels we focus on modeled exposures for school-age children (ages 5-18), 

8 noting that percentages of asthmatic school-age children estimated to experience exposures of 

9 concern are virtually indistinguishable from those for all children, and that patterns of exposure 

10 in children represent a broader range of at-risk populations, which includes adult asthmatics and 

11 older adults. 

12 In this section, we consider the following question: 

13 To what extent are potential alternative standards with revised levels estimated to 

14 reduce the occurrence of O3 exposures of concern, compared to the current 

15 standard, and what are the nature and magnitude of the exposures remaining 

16 for each alternative standard level evaluated? 

17 Key results related to this question are summarized below (Figures 4-1 to 4-4). Figures 4-1 and 

18 4-2 present estimates of one or more exposures of concern, and Figures 4-3 and 4-4 present 

19 estimates of two or more exposures of concern. 
4-21

1 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 

2 quality adjusted to just meet the current and potential alternative standards (averaged over 2006 to 2010) 

3

4
4-22

1 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 

2 quality adjusted to just meet the current and potential alternative standards (worst-case year from 2006 to 

201015 3 ) 

4

5

 

15“Worst-case” year refers to the year in each urban case study area with the largest percentage of children estimated to experience exposures of concern. 
4-23

1 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 

2 quality adjusted to just meet the current and potential alternative standards (averaged over 2006 to 2010) 

3

4

5
4-24

1 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 

2 quality adjusted to just meet the current and potential alternative standards (worst-case year from 2006 to 2010) 

3

4
4-25

1 As illustrated above in Figures 4-1 to 4-4, adjusting air quality to just meet progressively 

2 lower potential alternative standard levels reduces estimated exposures of concern consistently 

3 across urban case study areas. These results reflect the consistent reductions in the highest 

4 ambient O3 concentrations upon model adjustment, as summarized in section 3.2.1 and as 

5 discussed in more detail in the second draft HREA (U.S. EPA, 2014, chapter 4). Based on 

6 Figures 4-1 to 4-4 and the associated details described in the second draft HREA (U.S. EPA 

7 2014, chapter 5), we take note of the following with regard to exposures of concern for specific 

8 potential alternative standard levels: 

9 1. For a standard level of 70 ppb: 

10 a. On average over the years 2006 to 2010, a standard with a level of 70 ppb is 

11 estimated to allow approximately 3 to 11% of children in urban case study areas to 

12 experience one or more exposures of concern at or above 60 ppb (approximately 30 to 

13 70% reduction, relative to current standard). Summing across urban case study areas, 

14 these percentages correspond to over 1 million children experiencing over 1.5 million 

15 exposures of concern at or above 60 ppb during a single O3 season. Of these children, 

16 over 100,000 are asthmatics. 

17

18 b. On average over the years 2006 to 2010, a standard with a level of 70 ppb is 

19 estimated to allow approximately 0.5 to 3.5% of children in urban case study areas to 

20 experience two or more exposures of concern at or above 60 ppb (approximately 50 

21 to 85% reduction, relative to current standard). 

22

23 c. In the worst-case years (i.e., those with the largest exposure estimates), a standard 

24 with a level of 70 ppb is estimated to allow approximately 5 to 19% of children in 

25 urban case study areas to experience one or more exposures of concern at or above 60 

26 ppb, and approximately 2 to 9% to experience two or more. 

27

28 d. On average over the years 2006 to 2010, a standard with a level of 70 ppb is 

29 estimated to allow approximately 1% or less of children to experience one or more 

30 exposures of concern at or above 70 ppb (approximately 55 to 90% reduction, relative 

31 to current standard), and far less than 1% to experience two or more such exposures 

32 (approximately 65 to 100% reduction, relative to current standard). 

33

34 e. In the worst-case years, approximately 3% or less of children are estimated to 

35 experience one or more exposures of concern at or above 70 ppb, and less than 1% 

36 are estimated to experience two or more such exposures. 

37

38 f. A standard with a level of 70 ppb is estimated to allow less than 1% of children to 

39 experience one or more exposures of concern at or above 80 ppb, even in the worst40 case years. No children are estimated to experience two or more such exposures. 

41

42 2. For a standard level of 65 ppb: 
4-26

1 a. On average over the years 2006 to 2010, a standard with a level of 65 ppb is 

2 estimated to allow approximately 4% or less of children in urban case study areas to 

3 experience one or more exposures of concern at or above 60 ppb (approximately 70 to 

4 100% reduction, relative to current standard). Summing across urban case study 

5 areas, these percentages correspond to almost 500,000 children experiencing 

6 approximately 500,000 exposures of concern at or above 60 ppb during a single O3

7 season. Of these children, almost 50,000 are asthmatics. 

8

9 b. On average over the years 2006 to 2010, a standard with a level of 65 ppb is 

10 estimated to allow less than 1% of children to experience two or more exposures of 

11 concern at or above 60 ppb (approximately 85 to 100% reduction, relative to current 

12 standard). 

13

14 c. In the worst-case years, a standard with a level of 65 ppb is estimated to allow 

15 approximately 9% or less of children to experience one or more exposures of concern 

16 at or above 60 ppb, and approximately 3% or less to experience two or more such 

17 exposures. 

18

19 d. On average over the years 2006 to 2010, a standard with a level of 65 ppb is 

20 estimated to allow less than 1% of children to experience one or more exposures of 

21 concern at or above 70 ppb (approximately 90 to 100% reduction, relative to current 

22 standard), and no children to experience two or more such exposures (100% 

23 reduction, relative to current standard). Even in the worst-case years, a level of 65 

24 ppb is estimated to allow less than 1% of children to experience exposures of concern 

25 at or above 70 ppb. 

26

27 e. A standard with a level of 65 ppb is estimated to allow virtually no children to 

28 experience exposures of concern at or above 80 ppb, even in the worst-case years. 

29

30 3. For a standard level of 60 ppb: 

31 a. On average over the years 2006 to 2010, a standard with a level of 60 ppb is 

32 estimated to allow approximately 1% or less of children to experience one or more 

33 exposures of concern at or above 60 ppb (approximately 90 to 100% reduction, 

34 relative to current standard), and virtually no children to experience multiple such 

35 exposures. 

36

37 b. In the worst-case years, a standard with a level of 60 ppb is estimated to allow 

38 approximately 2% or less of children to experience one or more exposures of concern 

39 at or above 60 ppb, and virtually no children to experience multiple such exposures.

40

41 c. On average over the years 2006 to 2010, a standard with a level of 60 ppb is 

42 estimated to eliminate exposures of concern at or above 70 ppb (100% reduction, 

43 relative to current standard) or 80 ppb. Even in years with the highest exposure 

44 estimates, virtually no children are estimated to experience such exposures. 
4-27

1 In further considering these exposure estimates, we take note of the associated 

2 uncertainties, as discussed in more detail in section 3.2.2 of this second draft PA. These include 

3 (1) individual variability in responsiveness to O3 exposures; (2) potential to underestimate 

4 exposures in most highly exposed populations; and (3) potential to overestimate exposures in 

5 populations who alter behavior in response to high O3 days (i.e., spend less time being active 

6 outdoors). 

7 4.4.2.2 Risk-Based Considerations: Lung Function 

8 As discussed above in more detail in section 3.2.3.1 of this second draft PA, the 

9 assessment of lung function risks presented in the second draft HREA (U.S. EPA, 2014) 

10 provides estimates of the number and percent of people experiencing O3-induced lung function 

11 decrements greater than or equal to 10, 15, and 20%. In the last review, CASAC advised EPA to 

12 focus on decrements of 10% or greater when considering people with pre-existing lung disease 

13 (Samet, 2011). 

14 Lung function risk estimates are based on an updated dose-threshold model that estimates 

15 FEV1 responses for individuals following short-term exposures to O3 (McDonnell, Stewart, and 

16 Smith, 2010), reflecting methodological improvements since the last review (U.S. EPA, 2014, 

17 section 6.2.4). The approach taken in the second draft HREA to estimating O3-induced lung 

18 function decrements, and the key uncertainties associated with these estimates, are summarized 

19 in section 3.2.3.1 for air quality adjusted to just meet the current standard and are discussed in 

20 more detail in chapter 6 of the HREA (U.S. EPA, 2014). 

21 As discussed in section 3.2.3.1, in evaluating potential alternative standard levels we 

22 focus on modeled exposures for school-age children, with an emphasis on asthmatic children. As 

23 with exposures of concern, the percentages of all school age children and asthmatic school age 

24 children estimated to experience particular O3-induced lung function decrements are virtually 

25 indistinguishable. 

26 In this section, we consider the following question: 

27 To what extent are potential alternative standards with revised levels estimated to 

28 decrease the occurrence of O3-induced lung function decrements, compared to 

29 the current standard, and what are the nature and magnitude of the decrements 

30 remaining for each alternative standard level evaluated? 

31 Key results related to this question are summarized below (Figures 4-5 to 4-8). Figures 4-5 and 

32 4-6 present estimates of one or more O3-induced lung function decrements, and Figures 4-7 and 

33 4-8 present estimates of two or more decrements. 
4-28

1 Figure 4-5. Percent of children estimated to experience one or more O3-induced lung function decrements greater than 10, 15, 

2 or 20% for air quality adjusted to just meet the current and potential alternative standards (averaged over 2006 

3 to 2010) 

4

5
4-29

1 Figure 4-6. Percent of children estimated to experience one or more O3-induced lung function decrements greater than 10, 15, 

2 or 20% for air quality adjusted to just meet the current and potential alternative standards (worst-case year from 

3 2006 to 2010) 

4

5
4-30

1 Figure 4-7. Percent of children estimated to experience two or more O3-induced lung function decrements greater than 10, 

2 15, or 20% for air quality adjusted to just meet the current and potential alternative standards (averaged over 

3 2006 to 2010) 

4

5
4-31

1 Figure 4-8. Percent of children estimated to experience two or more O3-induced lung function decrements greater than 10, 

2 15, or 20% for air quality adjusted to just meet the current and potential alternative standards (worst-case year 

3 from 2006 to 2010) 

4

5
4-32

1 As illustrated above in Figures 4-5 to 4-8, adjusting air quality to just meet progressively 

2 lower potential alternative standard levels consistently reduces the percent of children estimated 

3 to experience potentially adverse lung function decrements. These results reflect the consistent 

4 reductions in the highest ambient O3 concentrations upon model adjustment (section 3.2.1; U.S. 

EPA, 2014, chapter 4).16 5 Based on Figures 4-5 to 4-8 and the associated details described in the 

6 second draft HREA (U.S. EPA 2014, chapter 6), we take note of the following with regard to 

7 specific potential alternative standard levels: 

8 1. For a standard level of 70 ppb: 

9 a. On average over the years 2006 to 2010, a standard with a level of 70 ppb is 

10 estimated to allow approximately 11 to 17% of children in urban case study areas, 

11 including asthmatic children, to experience one or more O3-induced lung function 

12 decrements ≥ 10% (approximately 6 to 27% reduction, relative to current 

13 standard) per season. Summing across case study areas, these percentages 

14 correspond to approximately 260,000 asthmatic children experiencing 

15 approximately 1 million total occurrences of O3-induced lung function 

16 decrements greater than or equal to 10%. 

17

18 b. On average over the years 2006 to 2010, a standard with a level of 70 ppb is 

19 estimated to allow approximately 6 to 11% of children, including asthmatic 

20 children, to experience two or more O3-induced lung function decrements ≥ 10% 

21 (approximately 8 to 30% reduction, relative to current standard). 

22

23 c. In the worst-case years, a standard with a level of 70 ppb is estimated to allow 

24 approximately 14 to 20% of children, including asthmatic children, to experience 

25 one or more O3-induced lung function decrements ≥10%, and approximately 7 to 

26 13% to experience two or more such decrements. 

27

28 d. On average over the years 2006 to 2010, a standard with a level of 70 ppb is 

29 estimated to allow approximately 2 to 4% of children, including asthmatic 

30 children, to experience one or more O3-induced lung function decrements ≥ 15%, 

31 and approximately 1 to 2.5% of children to experience two or more such O3-

32 induced decrements. In the worst-case years, approximately 3 to 5% of children 

33 are estimated to experience one or more O3-induced lung function decrements 

34 ≥15%, and approximately 1 to 3% are estimated to experience two or more such 

35 decrements. 

36

37 e. A standard with a level of 70 ppb is estimated to allow 2% or fewer children to 

38 experience any O3-induced lung function decrements ≥ 20%, even in the worst39 case years. Approximately 1% or fewer children are estimated to experience two 

 

16The impact of the dose threshold in the lung function risk model is that O3-induced FEV1 decrements result 

primarily from exposures to O3 concentrations above about 40 ppb (US EPA, 2013, chapter 6). 
4-33

1 or more O3-induced lung function decrements ≥ 20%, even in the worst-case 

2 years. 

3

4 2. For a standard level of 65 ppb: 

5 a. On average over the years 2006 to 2010, a standard with a level of 65 ppb is 

6 estimated to allow approximately 3 to 15% of children, including asthmatic 

7 children, to experience one or more O3-induced lung function decrements ≥ 10% 

8 (approximately 20 to 77% reduction, relative to current standard). Summing 

9 across urban case study areas, these percentages correspond to approximately 

10 190,000 asthmatic children experiencing almost 750,000 total occurrences of O3-

11 induced lung function decrements ≥ 10%. 

12

13 b. On average over the years 2006 to 2010, a standard with a level of 65 ppb is 

14 estimated to allow approximately 1 to 9% of children, including asthmatic 

15 children, to experience two or more O3-induced lung function decrements ≥ 10% 

16 (approximately 20 to 80% reduction, relative to current standard). 

17

18 c. In the worst-case years, a standard with a level of 65 ppb is estimated to allow 

19 approximately 4 to 18% of children to experience one or more O3-induced lung 

20 function decrements ≥ 10%, and approximately 2 to 11% to experience two or 

21 more such decrements. 

22

23 d. On average over the years 2006 to 2010, a standard with a level of 65 ppb is 

24 estimated to allow approximately 3% or less of children to experience one or 

25 more O3-induced lung function decrements ≥ 15%, and approximately 2% or less 

26 of children to experience two or more such O3-induced decrements. In the worst27 case years, approximately 4% or less of children are estimated to experience one 

28 or more O3-induced lung function decrements ≥ 15%, and up to approximately 

29 2.5% are estimated to experience two or more such decrements. 

30

31 e. A standard with a level of 65 ppb is estimated to allow less than 1.5% of children 

32 to experience any O3-induced lung function decrements ≥ 20%, even in the worst33 case years. A standard with a level of 65 ppb is estimated to allow less than 1% of 

34 children to experience two or more O3-induced lung function decrements ≥ 20%, 

35 even in the worst-case years. 

36

37 3. For a standard level of 60 ppb: 

38 a. On average over the years 2006 to 2010, a standard with a level of 60 ppb is 

39 estimated to allow approximately 5 to 11% of children, including asthmatic 

40 children, to experience one or more O3-induced lung function decrements ≥ 10% 

41 (approximately 35 to 77% reduction, relative to current standard). Summing 

42 across urban case study areas, these percentages correspond to approximately 

43 140,000 asthmatic children experiencing approximately 500,000 total occurrences 

44 of O3-induced lung function decrements ≥ 10%. 

45
4-34

1 b. On average over the years 2006 to 2010, a standard with a level of 60 ppb is 

2 estimated to allow approximately 2 to 6% of children to experience two or more 

3 O3-induced lung function decrements ≥ 10% (approximately 40 to 70% reduction, 

4 relative to current standard). 

5

6 c. In the worst-case years, a standard with a level of 60 ppb is estimated to allow 

7 approximately 5 to 13% of children to experience one or more O3-induced lung 

8 function decrements ≥ 10%, and approximately 2 to 7% to experience two or 

9 more such decrements. 

10

11 d. A standard with a level of 60 ppb is estimated to allow less than about 3% of 

12 children to experience any O3-induced lung function decrements ≥ 15% and less 

13 than 1% to experience decrements greater than 20%, even in years with the 

14 highest exposure estimates. A standard with a level of 60 ppb is estimated to 

15 allow less than 1.5% of children to experience two or more O3-induced lung 

16 function decrements ≥ 15% and less than 0.5% to experience two or more 

17 decrements ≥ 20%, even in years with the highest exposure estimates. 

18 In further considering these exposure estimates, we take note of the associated 

19 uncertainties, as discussed in more detail in section 3.2.2 of this second draft PA. In addition to 

20 the uncertainties in exposure estimates noted above, these include the relative lack of exposure21 response information for key at-risk populations (i.e., children and asthmatics), since most 

22 controlled human exposures studies are conducted in healthy adults. 

23 4.4.2.3 Risk-Based Considerations: Epidemiology-Based Mortality and Morbidity 

24 The epidemiology-based risk assessments presented in the second draft HREA (U.S. 

25 EPA, 2014, chapter 7) provide estimates of total mortality, respiratory hospital admissions and 

26 emergency department visits, and asthma exacerbations associated with short-term O3

27 concentrations. The HREA also presents estimates of respiratory mortality associated with longterm17 28 concentrations. In evaluating these risk estimates, we consider the following question: 

29 To what extent are potential alternative standards with revised levels estimated to 

30 decrease O3 health risks, compared to the current standard, and what are the 

31 nature and magnitude of the health risks remaining for each alternative standard 

32 level evaluated? 

33 As discussed in more detail in section 3.2.3.2 of this second draft PA, in considering this 

34 question we are mindful that the model-based approach used to adjust air quality in the second 

 

17Estimates 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. 
4-35

1 draft HREA has important implications for risk estimates developed by applying concentration2 response relationships from epidemiologic studies (section 3.2.1). In particular, we note the 

3 uncertainty associated with using such concentration-response relationships to estimate risks for 

4 model-adjusted air quality with spatial and temporal patterns of ambient O3 that are different 

5 from those present in the epidemiologic study locations. In addition, given the use of linear 

6 concentration-response relationships, risk estimates are equally influenced by decreasing high O3

7 concentrations and increasing low O3 concentrations following model adjustment, when the 

8 increases and decreases are of equal magnitude. These and other uncertainties associated with 

9 risk estimates are discussed in section 3.2.3.2. 

10 Key results from the second draft HREA (U.S. EPA, 2014, chapter 7) are summarized 

11 below for estimates of total mortality associated with short-term O3 concentrations (Figures 4-9

12 and 4-10), respiratory hospital admissions associated with short-term O3 concentrations (Figure 

13 4-11), and respiratory mortality associated with long-term O3 concentrations (Figure 4-12). The 

14 other morbidity effects evaluated in the second draft HREA (i.e., respiratory emergency 

15 department visits and asthma symptoms associated with short-term concentrations) exhibit 

16 patterns across standard levels that are similar to those reported for total mortality and respiratory 

17 hospital admissions (U.S. EPA, 2014, chapter 7). 

18 As discussed in section 3.2.3.2, for total mortality associated with short-term O3

19 concentrations we consider estimates of risk based on the full distributions of area-wide O3

20 concentrations (Figure 4-9) and estimates of risk associated with various portions of those 

distributions (Figure 4-10).18 21 In doing so, we recognize the reduced certainty in a linear 

22 concentration-response relationship at the lower ends of air quality distributions, and the greater 

23 certainty in increased incidence and severity of effects at higher exposure concentrations 

(discussed in more detail in section 3.2.3.2).19 24

 

18The second draft HREA does not present distributions of risk over distributions of area-wide concentrations for 

other epidemiology-based risk endpoints (U.S. EPA, 2014, chapter 7). 

19As discussed in section 3.1.2.2, 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. 
4-36

1 Figure 4-9. Estimates of Total Mortality Associated with Short-Term O3 Concentrations in Urban Case Study Areas (Air 

2 Quality Adjusted to Current and Potential alternative standard levels) – Total Risk 

3

4
4-37

1 Figure 4-10. Estimates of Total Mortality Attributable to Days with 8-Hour Area-Wide O3 Concentrations at or above 20, 40, 

2 or 60 ppb, Summed Across Urban Case Study Areas (Air Quality Adjusted to Current and Potential alternative 

3 standard levels) 

4

5

6
4-38

1 Figure 4-11. Estimates of Respiratory Hospital Admissions Associated with Short-Term O3 Concentrations in Urban Case 

2 Study Areas (Air Quality Adjusted to Current and Potential alternative standard levels) – Total Risk 

3

4
4-39

1 Figure 4-12. Estimates of Respiratory Mortality Associated with long-term O3 Concentrations in Urban Case Study Areas (Air 

2 Quality Adjusted to Current and Potential alternative standard levels) – Total Risk 

3
4-40

1 Based on Figures 4-9 to 4-12 and the associated details described in the second draft 

2 HREA (U.S. EPA 2014, chapter 7), we take note of the following for a standard level of 70 ppb: 

3 1. Total mortality associated with short-term O3 concentrations: 

4 a. A standard level of 70 ppb results in modest changes in total risk, compared to the 

5 current standard. Across urban case study areas, risks are estimated to decrease by 

6 up to approximately 5%. These risk reductions are estimated most consistently for 

7 the model year with generally higher O3-associated risks (2007). In the year with 

8 generally lower risks (2009), a standard level of 70 ppb results in either no change 

9 or small reductions in estimated risks in most urban case study areas. In one area 

10 (Detroit) for the 2009 model year, O3-associated mortality is estimated to increase 

11 by approximately 4%, compared to the current standard (see section 3.2.3.2 for 

further discussion of increased risk estimates following model adjustment20 12 ). 

13

14 b. When summed across urban case study areas, a standard level of 70 ppb is 

15 estimated to reduce O3-associated total mortality by approximately 4% (2007 

16 model year) and 2% (2009 model year), compared to the current standard. For 

17 days with area-wide concentrations at or above 40 ppb, a standard level of 70 ppb 

18 is estimated to reduce O3-associated total mortality by approximately 9% (2007 

19 model year) and 8% (2009 model year). For days with area-wide concentrations at 

20 or above 60 ppb, a standard level of 70 ppb is estimated to reduce O3-associated 

21 total mortality by approximately 50% (2007 model year) and 70% (2009 model 

year).21 22

23

24 2. Respiratory hospital admissions associated with short-term O3 concentrations: Compared to 

25 the current standard, changes in total risk estimated for a standard level of 70 ppb are similar 

26 to the changes in total risks estimated for total mortality (U.S. EPA, 2014, chapter 7). 

27

28 3. Respiratory mortality associated with long-term O3 concentrations: A standard level of 70 

29 ppb reduces total risk, compared to the current standard. Across urban case study areas, risks 

30 are estimated to decrease by up to approximately 8%. These risk reductions are estimated 

31 most consistently for the model year with generally higher O3-associated risks (2007). In the 

32 year with generally lower O3 concentrations (2009), a standard level of 70 ppb results in 

33 either no change or small reductions in estimated risks in most urban case study areas. In one 

34 area (Detroit) for the 2009 model year, O3-associated mortality is estimated to increase by 

35 approximately 1%, compared to the current standard.

 

20As 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. 

21These results reflect the fact that increases in area-wide O3 concentrations upon model adjustment occur primarily 

at relatively low concentrations (i.e., at or below area-wide concentrations of approximately 45 ppb) (U.S. EPA, 

2014, section 4.3.3.2 and appendix 7B). 
4-41

1 Based on Figures 4-9 to 4-12 and the associated details described in the second draft 

2 HREA (U.S. EPA 2014, chapter 7), we take note of the following for a standard level of 65 ppb: 

3 1. Total mortality associated with short-term O3 concentrations: 

4 a. A standard level of 65 ppb results in small changes in total risk, compared to the 

5 current standard. Across most urban case study areas, risks are estimated to 

6 decrease by up to approximately 9%. In one area (New York City), risks are 

estimated to decrease by up to approximately 22%.22 7 These risk reductions are 

8 estimated most consistently for the model year with generally higher O3-

9 associated risks (2007). In the year with generally lower risks (2009), a standard 

10 level of 65 ppb results in smaller reductions in estimated risks in most urban case 

11 study areas. In one area (Detroit) for the 2009 model year, O3-associated mortality 

12 is estimated to increase by approximately 1% compared to the current standard. 

13

14 b. When summed across urban case study areas, a standard level of 65 ppb is 

15 estimated to reduce O3-associated total mortality by approximately 11% (2007 

16 model year) and 8% (2009 model year), compared to the current standard. For 

17 days with area-wide concentrations at or above 40 ppb, a standard level of 65 ppb 

18 is estimated to reduce O3-associated total mortality by almost 40% (2007 and 

19 2009 model years). For days with area-wide concentrations at or above 60 ppb, a 

20 standard level of 65 ppb is estimated to reduce O3-associated total mortality by 

21 over 80% (2007 and 2009 model years). 

22

23 2. Respiratory hospital admissions associated with short-term O3 concentrations: Compared to 

24 the current standard, changes in total risk estimated for a standard level of 65 ppb are similar 

25 to the changes in total risk estimated for total mortality (U.S. EPA, 2014, chapter 7). 

26

27 3. Respiratory mortality associated with long-term O3 concentrations: A standard level of 65

28 ppb reduces total risk, compared to the current standard. Across most urban case study areas, 

29 risks are estimated to decrease by up to approximately 10%. In one area (New York City), 

30 risks are estimated to decrease by up to approximately 22%. Risk reductions are estimated 

31 across all urban case study areas and in both model years evaluated, with larger reductions 

32 estimated for 2007 (i.e., the model year with generally higher O3-associated risks). 

33 Based on Figures 4-9 to 4-12 and the associated details described in the second draft 

34 HREA (U.S. EPA 2014, chapter 7), we take note of the following for a standard level of 60 ppb: 

35 1. Total mortality associated with short-term O3 concentrations: 

36 a. A standard level of 60 ppb is estimated to reduce total risk, compared to the 

37 current standard, in all urban case study areas. Across urban case study areas, 

 

22As discussed in the second draft HREA (U.S. EPA, 2014, section 4.5), the New York and Los Angeles urban case 

study areas required the largest reductions in NOX in order to meet the existing and potential alternative standards. 

The HDDM-based O3 estimates become more uncertain for larger changes in precursor emissions, and the HREA 

notes less overall confidence in results for New York and Los Angeles.
4-42

1 risks are estimated to decrease by up to approximately 15%. Estimated risk 

2 reductions are larger for the model year with generally higher O3-associated risks 

3 (2007). 

4

5 b. When summed across urban case study areas, a standard level of 60 ppb is 

6 estimated to reduce O3-associated total mortality by approximately 14% (2007 

7 model year) and 10% (2009 model year), compared to the current standard. For 

8 days with area-wide concentrations at or above 40 ppb, a standard level of 60 ppb 

9 is estimated to reduce O3-associated total mortality by approximately 50% (2007 

10 and 2009 model years). For days with area-wide concentrations at or above 60 

11 ppb, a standard level of 60 ppb is estimated to reduce O3-associated mortality by 

12 over 95% (2007 and 2009 model years). 

13

14 2. Respiratory hospital admissions associated with short-term O3 concentrations: Compared to 

15 the current standard, changes in total risk estimated for a standard level of 60 ppb are similar 

16 to the changes in total risk estimated for total mortality (U.S. EPA, 2014, chapter 7). 

17

18 3. Respiratory mortality associated with long-term O3 concentrations: A standard level of 60

19 ppb reduces total risk, compared to the current standard. Across urban case study areas, risks 

20 are estimated to decrease by up to approximately 17%. Risk reductions are estimated across 

21 all urban case study areas and in both model years evaluated, with larger reductions 

22 estimated for 2007 (i.e., the model year with generally higher O3-associated risks).

23 In further considering these risk estimates, we take note of the associated uncertainties, as 

24 discussed in more detail in section 3.2.3.2 of this second draft PA. These include (1) the national 

25 representativeness of urban case study areas in terms of the O3 response to reductions in NOX

26 emissions; (2) the shape of the concentration-response function at lower ambient concentrations; 

27 (3) the use of concentration-response relationships developed for particular populations in 

28 particular locations to estimate health risks in different populations and locations; (4) the 

29 applications of concentration-response relationships to model-adjusted air quality, given the 

30 altered spatial/temporal patterns of ambient O3 and the potential for increases in relatively low 

31 O3 concentrations to increase risk estimates; and (5) the possibility for reductions in risk 

32 associated with reductions in PM and/or NO2 resulting from control of NOX. 

33 4.5 CASAC ADVICE 

34 In the fall of 2011, rather than revising the O3 NAAQS as part of the reconsideration 

35 process, EPA coordinated further proceedings on the reconsideration rulemaking with the current 

36 ongoing periodic review. Accordingly, in this section we are briefly describing CASAC advice 

37 from the reconsideration of the 2008 final decision as well as CASAC advice received during the 

38 current review as it pertains to potential alternative standards. 
4-43

1 Consistent with their advice in 2008, CASAC reiterated during the reconsideration its 

2 support for an 8-hour primary O3 standard with a level ranging from 60 to 70 ppb, combined 

3 with the current form. Specifically, in response to EPA’s solicitation of their advice during the 

4 reconsideration, the CASAC letter (Samet 2010) to the Administrator stated: 

5 CASAC fully supports EPA’s proposed range of 0.060 – 0.070 parts per million 

6 (ppm) for the 8-hour primary ozone standard. CASAC considers this range to be 

7 justified by the scientific evidence as presented in the Air Quality Criteria for 

8 Ozone and Related Photochemical Oxidants (March 2006) and Review of the 

9 National Ambient Air Quality Standards for Ozone: Policy Assessment of 

10 Scientific and Technical Information, OAQPS Staff Paper (July 2007). 

11

12 Similarly, in response to EPA’s request for additional advice on the reconsideration in 

13 2011, CASAC reaffirmed their conclusion that “the evidence from controlled human and 

14 epidemiological studies strongly supports the selection of a new primary ozone standard within 

15 the 60 – 70 ppb range for an 8-hour averaging time” (Samet, 2011). CASAC further concluded 

16 that this range “would provide little margin of safety at its upper end” (Samet, 2011, p. 2). 

17 In the first draft PA, staff concluded that the available evidence provides support for 

18 conducting further exposure and risk analyses of potential alternative standard levels in the range 

19 of 60 to 70 ppb (USEPA, 2012b). In response, CASAC noted that the draft PA provided “a 

20 strong scientific rationale for consideration of ozone levels (8 hour averages) of 60 ppb to 70 

21 ppb” (Frey and Samet, 2012). 

22 4.6 PRELIMINARY STAFF CONCLUSIONS ON ALTERNATIVE PRIMARY 

23 STANDARDS FOR CONSIDERATION 

24 Staff’s consideration of alternative primary O3 standards builds upon our conclusion, 

25 discussed in section 3.4, that the overall body of evidence and exposure/risk information calls 

26 into question the adequacy of public health protection afforded by the current standard, 

27 particularly for at-risk populations. In section 3.4, we further conclude that it is appropriate in 

28 this review to consider alternative standards that would increase public health protection, 

29 compared to the current standard, and that it is not appropriate to consider alternative standards 

30 with levels higher than the current standard, which would decrease public health protection. 

31 As an initial matter, for the reasons discussed in section 4.1 above, we conclude it is 

32 appropriate to continue using O3 as the indicator for the standard that protects against exposures 

33 to ambient O3 and other photochemical oxidants. For the reasons discussed in sections 4.2 and 

34 4.3 above, we also conclude that it is appropriate for the Administrator to consider retaining the 

35 current averaging time and form. In the remainder of this section, we present our more focused 

36 discussion on the range of alternative levels that, in our judgment, it is appropriate for the 

37 Administrator to consider. 
4-44

1 For a standard that is defined in terms of the current indicator, averaging time, and form, 

2 we reach the conclusion that, depending on the public health policy judgments made by the 

3 Administrator, the evidence and information available in this review supports consideration of 

alternative levels from 70 down to 60 ppb (section 4.4, above).23 4 Compared to the current 

5 standard, a revised standard with a level from 70 to 60 ppb would be expected to increase public 

6 health protection against both short- and long-term O3 exposures, including for members of at7 risk populations. The scientific evidence and exposure/risk estimates that could support revised 

8 standards with levels from the upper, middle, and lower portions of this range are summarized 

9 below, with a specific focus on levels of 70 ppb, 65 ppb, and 60 ppb. Key exposure/risk 

10 information is summarized in Tables 4-4 and 4-5, and Figure 4-13. 

11

 

23As discussed in sections 3.1.2, 3.2, and 3.4 of this second draft PA, we further conclude that it would not be 

appropriate to consider a standard level higher than 75 ppb, which would decrease public health protection 

compared to the current standard. 
4-45

1 Table 4-4. Summary of Estimated Exposures of Concern for Potential alternative 

standard levels of 70, 65, and 60 ppb in Urban Case Study Areas24 2

Benchmark 

Level

Alternative 

Standard 

Level (ppb)

Average % 

Children

Exposed25

Number of Children (5 

to 18 years)

[Number of Asthmatic 

Children]26

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 95,000 [10,000] 73 3.2

65 0-0.2 16,000 [1,600] 95 0.5

60 0 0 [0] 100 0.1

≥ 60 ppb

70 3.3-10.2 1,177,000 [122,000] 46 18.9

65 0-4.2 392,000 [40,000] 80 9.5

60 0-1.2 69,000 [7,000] 96 2.2

Two or more exposures of concern per season

≥ 70 ppb

70 0-0.1 3,000 [360] 95 0.4

65 0 0 [0] 100 0

60 0 0 [0] 100 0

≥ 60 ppb

70 0.5-3.5 319,000 [33,000] 61 9.2

65 0-0.8 65,000 [7,000] 92 2.8

60 0-0.2 3,800 [400] 100 0.3

3

4

 

24As 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. 

25Estimates for each urban case study area were averaged for the years evaluated in the second draft HREA (2006 to 

2010). Ranges reflect the ranges across urban case study areas. 

26Numbers 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. 
4-46

1 Table 4-5. Summary of Estimated Lung Function Decrements for Potential alternative 

2 standard levels of 70, 65, and 60 ppb in Urban Case Study Areas 

Lung 

Function 

Decrement

Alternative 

Standard 

Level

Average % 

Children27

Number of Children (5 

to 18 years) [Number of 

Asthmatic Children]28

Average % 

Reduction from 

Current Standard

% Children 

Worst Year and 

Area

One or more decrements per season

≥ 10%

70 11-17 2,547,000 [263,000] 15 20

65 3-15 1,931,000 [195,000] 31 18

60 5-11 1,392,000 [139,000] 45 13

≥ 15%

70 2-4 564,000 [58,000] 26 5

65 0-3 355,000 [36,000] 50 4

60 1-2 224,000 [22,000] 67 3

≥ 20%

70 1-2 189,000 [20,000] 32 2.1

65 0-1 107,000 [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 1,418,000 [146,000] 17 13

65 1.3-8.8 1,020,000 [102,000] 37 11

60 2.1-6.4 743,000 [74,000] 51 7.3

≥ 15%

70 0.9-2.4 274,000 [28,000] 29 3.1

65 0.1-1.8 169,000 [17,000] 54 2.3

60 0.2-1.0 100,000 [10,000] 71 1.4

≥ 20%

70 0.3-0.8 84,000 [9,000] 34 1.1

65 0-0.5 40,000 [4,000] 66 0.8

60 0-0.2 22,000 [2,000] 83 0.4

3

4

 

27Estimates in each urban case study area were averaged for the years evaluated in the second draft HREA (2006 to 

2010). Ranges reflect the ranges across urban case study areas. 

28Numbers 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. 
4-47

1 Figure 4-13. Estimates of Total Mortality Attributable to Days with 8-Hour Area-Wide O3

2 Concentrations at or above 20, 40, or 60 ppb - Risks Summed Across Urban 

3 Case Study Areas and Expressed Relative to 75 ppb 

4

5
4-48

1 As an initial observation, we note that controlled human exposure studies provide the 

2 most certain evidence indicating the occurrence of health effects in humans following exposures 

3 to specific O3 concentrations. As discussed above and in section 3.1.2.1, controlled human 

4 exposure studies have reported a variety of respiratory effects in healthy adults following 

5 exposures to O3 concentrations of 60, 70, or 80 ppb, and higher. The largest respiratory effects, 

6 and the broadest range of effects, have been reported following exposures to 80 ppb O3 or higher, 

7 in part because most exposure studies have been conducted at these higher concentrations. 

8 Exposures to O3 concentrations of 80 ppb or higher have been reported to decrease lung function, 

9 increase airway inflammation, increase respiratory symptoms, result in airway 

10 hyperresponsiveness, and decrease lung host defenses in healthy adults. Most of these effects 

have also been reported in healthy adults following exposures to lower O3 concentrations.29 11

12 Exposures to O3 concentrations of 70 ppb have been reported to decrease lung function and 

13 increase respiratory symptoms, a combination that meets the ATS criteria for an “adverse” 

14 response (section 3.1.3). Exposures to O3 concentrations of 60 ppb have been demonstrated to 

15 decrease lung function, with decrements in some individuals large enough to be judged an 

16 abnormal response by ATS, and which could be adverse in individuals with lung disease.

17 Exposures to O3 concentrations of 60 ppb have also been reported in one study to increase 

18 airway inflammation, which provides a mechanism by which O3 may cause other more serious 

19 respiratory effects (e.g., asthma exacerbations). 

20 Compared to the current standard, the second draft HREA estimates that a revised 

21 standard with a level of 70 ppb would reduce exposures of concern to O3 concentrations of 60, 

22 70, and 80 ppb, with such a standard level estimated to be most effective at limiting exposures at 

23 or above the higher health benchmark concentrations. On average over the years 2006 to 2010, a 

24 standard with a level of 70 ppb is estimated to allow only up to about 1% of children (i.e., ages 5 

25 to 18) to experience exposures of concern at or above 70 ppb (73% reduction, compared to 

26 current standard), and far less than 1% to experience two or more such exposures (95% 

27 reduction, compared to current standard). In the worst-case location and year (i.e., location and 

28 year with the largest exposure estimate), about 3% of children are estimated to experience one or 

29 more exposures of concern at or above 70 ppb, and less than 1% are estimated to experience two 

30 or more. A standard with a level of 70 ppb is estimated to allow far less than 1% of children to 

31 experience exposures of concern at or above the 80 ppb benchmark concentration, even in the 

32 worst-case year (Table 4-4).

 

29Though airway hyperresponsiveness and 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. 
4-49

1 A standard with a level of 70 ppb is estimated to allow about 3 to 10% of children to 

2 experience one or more exposures of concern at or above 60 ppb in a single O3 season. 

3 Compared to the current standard, this reflects about a 46% reduction, on average across urban 

4 case study areas. A standard with a level of 70 ppb is estimated to allow about 1% to 4% of 

5 children to experience two or more exposures of concern at or above 60 ppb. In the worst-case 

6 location and year, a standard set at 70 ppb is estimated to allow about 19% of children to 

7 experience one or more exposures of concern at or above 60 ppb, and 9% to experience two or 

8 more such exposures (Table 4-4). 

9 Compared to the current standard, the second draft HREA estimates that a revised 

10 standard with a level of 70 ppb would also reduce O3-induced lung function decrements in 

11 children. A level of 70 ppb is estimated to be most effective at limiting the occurrences of 

12 moderate and large lung function decrements (i.e., FEV1 decrements ≥ 15% and ≥ 20%, 

13 respectively). On average over the years 2006 to 2010, a standard with a level of 70 ppb is 

14 estimated to allow about 2 to 4% of children to experience one or more moderate O3-induced 

15 lung function decrements (i.e., ≥ 15%), which would be of concern for healthy people, and about 

16 1 to 2.5% of children to experience two or more such decrements (approximately 30% reduction, 

17 compared to the current standard). In the worst-case location and year, up to 5% of children are 

18 estimated to experience one or more O3-induced lung function decrements ≥ 15%, and up to 3% 

19 are estimated to experience two or more such decrements. A standard set at 70 ppb is estimated 

20 to allow about 2% or fewer children to experience large O3-induced lung function decrements 

21 (i.e., ≥ 20%), and to allow about 1% or fewer children to experience two or more such 

22 decrements, even in the years and locations with the largest exposure estimates (Table 4-5). 

23 On average over the years 2006 to 2010, a standard set at 70 ppb is estimated to allow 

24 about 11 to 17% of children to experience one or more moderate O3-induced lung function 

25 decrements (i.e., ≥ 10%), which could be adverse for people with lung disease. This reflects an 

26 average reduction of about 15%, compared to current standard. A standard with a level of 70 ppb 

27 is also estimated to allow about 6 to 11% of children to experience two or more such decrements 

28 (17% reduction, compared to current standard). In the worst-case location and year, a standard 

29 set at 70 ppb is estimated to allow about 20% of children, including asthmatic children, to 

30 experience one or more O3-induced lung function decrements ≥ 10%, and 13% to experience two 

31 or more such decrements (Table 4-5). 

32 With regard to our analyses of epidemiologic studies we note that, compared to the 

33 current standard, a revised standard with a level of 70 ppb would be more effective in 

34 maintaining short-term ambient O3 concentrations below those present in locations that provided 

35 the basis for positive and statistically significant health effect associations. In particular, the 

36 study by Mar and Koenig (2009) reported positive and statistically significant associations with 
4-50

1 respiratory emergency department visits in Seattle, a location that met the current standard over 

2 the entire study period but that would have violated a revised standard with a level of 70 ppb. In 

3 addition, most of the Canadian study cities evaluated by Dales et al. (2006), who reported 

4 positive and statistically significant associations with respiratory hospital admissions, would 

5 have met the current standard over the entire study period but would have violated a revised 

6 standard with a level of 70 ppb over at least part of the study period (Table 4-1). Finally, 

7 compared to the current standard, fewer of the study cities evaluated by Katsouyanni et al. 

8 (2009) (for mortality and hospital admissions) and Bell et al. (for the 30 ppb cut-point) would 

have met a revised standard with a level of 70 ppb (Tables 4-1, 4-2).30 9

10 With regard to long-term O3 concentrations, air quality analyses indicate that in 9 out of 

11 the 12 urban case study areas a revised standard with a level of 70 ppb would be expected to 

12 maintain long-term O3 concentrations below those where the study by Jerrett et al. (2009) 

13 indicates the most confidence in the reported association with respiratory mortality. This is 

14 compared to 6 out of 12 areas for the current standard. 

15 In further considering the potential implications of epidemiology studies for potential 

16 alternative standard levels, we note that the emphasis given to estimates of O3-associated 

17 mortality or morbidity risks will depend in large part on the extent to which key uncertainties in 

18 these estimates are emphasized (e.g., uncertainties in applying concentration-response 

19 relationships from epidemiologic studies to adjusted air quality with different spatial/temporal 

20 distributions of ambient O3). To the extent emphasis is placed on estimates of O3-associated 

21 mortality and morbidity risks, it could be judged appropriate to focus on estimates of total risk 

22 (i.e., based on the full distributions of ambient O3 concentrations) and/or on estimates of risk 

23 associated with O3 concentrations in the upper portions of ambient distributions. 

24 A focus on total risks could be judged appropriate to the extent greater emphasis is placed 

25 on the possibility that concentration-response relationships remain linear over the entire 

26 distributions of ambient O3 concentrations, and thus to the extent greater emphasis is placed on 

27 the potential for mortality and/or morbidity to be affected by changes in relatively low O3

28 concentrations (as discussed above and in section 3.2.3.2). When summed across urban case 

29 study areas, a standard with a level of 70 ppb is estimated to reduce total mortality associated 

30 with short-term O3 exposures by about 2 to 3%, compared to the current standard (with 

reductions up to about 5% for individual urban case study areas).31 31 Based on a national modeling 

32 analysis, the majority of the U.S. population would be expected to experience modest reductions 

33 in such risks upon reducing precursor emissions. A standard with a level of 70 ppb is estimated 

 

30Though in the analyses presented in both of these studies, the majority of cities evaluated would have met a 

standard with a level of 70 ppb over the entire study periods. 

31Similar changes are estimated for respiratory morbidity associated with short-term O3 concentrations. 
4-51

1 to reduce respiratory mortality associated with long-term O3 concentrations by up to about 8% in 

2 urban case study areas. For both short- and long- term metrics, risk reductions are larger, and are 

3 estimated more consistently across urban case study areas, in a year with relatively higher O3

4 concentrations. 

5 A focus on risks associated with O3 concentrations in the upper portions of ambient 

6 distributions may reasonably be judged appropriate in light of the overall evidence indicating 

7 increasing magnitude, severity, and incidence of O3-attributable effects as exposure 

8 concentrations increase, as well as the greater uncertainty associated with the shapes of 

9 concentration-response curves for O3 concentrations in the lower portions of ambient 

10 distributions (section 3.2.3.2 and 3.4). There is no single ambient concentration below which 

11 uncertainty in O3-attributable effects increases notably in all locations, for all health endpoints, 

12 and in all populations. Therefore, we consider the distribution of mortality associated with 

13 various portions of the distribution of area-wide O3 concentrations (Figure 4-13, above). 

14 For days with area-wide concentrations at or above 20 ppb, a standard with a level of 70

15 ppb is estimated to reduce mortality associated with short-term O3 exposures by about 2 to 3% 

16 compared to the current standard, when O3-associated deaths are summed across urban case 

17 study areas. For days with area-wide concentrations at or above 40 ppb, a standard with a level 

18 of 70 ppb is estimated to reduce mortality associated with short-term O3 exposures by about 8 to 

19 9% compared to the current standard. For days with area-wide concentrations at or above 60 ppb, 

20 a standard with a level of 70 ppb is estimated to reduce O3-associated mortality by about 50% to 

70% (Figure 4-13).32 21

22 Based on all of the above considerations, we conclude that a standard with a level of 70 

23 ppb would be expected to provide additional incremental protection over that provided by the 

24 current O3 standard. A level of 70 ppb could be supported to the extent more emphasis is placed 

25 on the public health importance of higher O3 exposure concentrations (e.g., > 70, 80 ppb), larger 

26 O3-induced lung function decrements (e.g., > 15, 20%), and estimates of multiple occurrences of 

27 exposures of concern and O3-induced lung function decrements. In addition, a revised standard 

28 with a level of 70 ppb would be expected to be more effective than the current standard at 

29 maintaining short- and long-term ambient O3 concentrations below those where we have the 

30 most confidence in associations with mortality and morbidity. Overall across the U.S., such a 

 

32Fewer than 10% of total O3-associated deaths are attributable to days with such high area-wide concentrations, due 

to the relatively small number of days with area-wide concentrations of 60 ppb or above. 
4-52

1 standard would also be expected to reduce risks of O3-associated mortality and morbidity, 

particularly the portions of the risk associated with relatively high ambient concentrations.33 2

3 Next, we consider a standard with a level of 65 ppb. A level of 65 ppb is well-below the 

4 lowest O3 exposure concentration (i.e., 80 ppb) that has been reported to elicit a wide range of 

5 effects, including: lung function decrements, airway inflammation, respiratory symptoms, airway 

6 hyperresponsiveness, and decreased lung host defense in healthy adults, as noted above. A 

7 standard level of 65 ppb is also somewhat below the lowest exposure concentration at which the 

8 combined occurrence of respiratory symptoms and lung function decrements has been reported 

9 (i.e., 70 ppb), a combination judged adverse by the ATS (section 3.1.3). A level of 65 ppb is 

10 above the lowest exposure concentration demonstrated to result in lung function decrements 

11 large enough to be judged an abnormal response by ATS, and that could be adverse in 

12 individuals with lung disease. A level of 65 ppb is also above the lowest exposure concentration 

13 at which pulmonary inflammation has been reported (i.e., 60 ppb). 

14 Compared to the current standard and a revised standard with a level of 70 ppb, the 

15 second draft HREA estimates that a standard with a level of 65 ppb would reduce exposures of 

16 concern to the range of O3 benchmark concentrations analyzed (i.e., 60, 70, and 80 ppb). The 

17 HREA estimates that meeting a standard with a level of 65 ppb would eliminate exposures of 

18 concern at or above 80 ppb in urban case study areas. Such a standard is estimated to allow far 

19 less than 1% of children to experience one or more exposures of concern at or above the 70 ppb 

20 benchmark level, even in the years and locations with the largest exposure estimates, and is 

21 estimated to eliminate the occurrence of two or more exposures at or above 70 ppb (Table 4-4). 

22 In addition, on average over the years 2006 to 2010, a standard with a level of 65 ppb is 

23 estimated to allow between 0 and about 4% of children in urban case study areas to experience 

24 exposures of concern at or above 60 ppb. This reflects an 80% reduction (on average across 

25 areas), relative to the current standard. A standard with a level of 65 ppb is estimated to allow 

26 less than 1% of children to experience two or more exposures of concern at or above 60 ppb (> 

27 90% reduction, compared to current standard). In the worst-case location and year, about 10% of 

28 children are estimated to experience one or more exposures of concern at or above 60 ppb, with 

29 about 3% estimated to experience two or more such exposures (Table 4-4). 

30 Compared to the current standard and a revised standard with a level of 70 ppb, the 

31 second draft HREA estimates that a standard with a level of 65 ppb would also reduce the 

 

33In reaching this conclusion we recognize that, as discussed in detail in chapter 3 (sections 3.2.1 and 3.2.3.2), 

reducing NOX emissions to meet alternative O3 standards could result in increases in relatively low ambient 

concentrations. When we consider the epidemiologic-based estimates in light of all of the health effects evidence, 

and the considerations discussed more fully in section 3.2.3.2 above, we have greater certainty in the health benefits 

of reducing high ozone concentrations, and appreciable uncertainty regarding estimates of risk at lower 

concentrations. Accordingly, we judge that the range of levels discussed here is appropriate. 
4-53

1 occurrence of O3-induced lung function decrements. A level of 65 ppb is estimated to allow 

2 about 4% or less of children to experience O3-induced lung function decrements ≥ 15% (50% 

3 reduction, compared to current standard), even considering the worst-case location and year.

4 Such a standard is estimated to allow about 2% or less of children to experience two or more 

5 such decrements. A standard set at 65 ppb is estimated to allow about 1% or less of children to 

6 experience large O3-induced lung function decrements (i.e., ≥ 20%), even in the worst-case year 

7 and location (Table 4-5). 

8 On average over the years 2006 to 2010, a standard with a level of 65 ppb is estimated to 

9 allow about 3 to 15% of children to experience one or more moderate O3-induced lung function 

10 decrements (i.e., ≥ 10%), which could be adverse for people with lung disease. This reflects an 

11 average reduction of about 30%, relative to current standard. A standard with a level of 65 ppb is 

12 also estimated to allow about 1 to 9% of children to experience two or more such decrements 

13 (37% reduction, compared to current standard). In the worst-case location and year, a standard 

14 set at 65 ppb is estimated to allow up to about 18% of children to experience one or more O3-

15 induced lung function decrements ≥ 10%, and up to 11% to experience two or more such 

16 decrements (Table 4-5). 

17 With regard to O3 epidemiologic studies we note that, compared to the current standard 

18 and a standard with a level of 70 ppb, a revised standard with a level of 65 ppb would be more 

19 effective in maintaining short-term ambient O3 concentrations below those present in locations 

20 that provided the basis for positive and statistically significant health effect associations. In 

21 addition to the studies by Mar and Koenig (2009) and Dales et al. (2006) (discussed above for a 

22 level of 70 ppb), a revised standard with a level of 65 ppb would not allow the ambient O3

23 concentrations that provided the basis for mortality associations in most of the Canadian study 

cities evaluated by Katsouyanni et al. (2009) (Table 4-1).34 24 In addition, fewer of the study cities 

25 evaluated by Bell et al. (for the 30 ppb cut-point) would have met a revised standard with a level 

of 65 ppb (Table 4-2).35 26

27 With regard to long-term O3 concentrations, air quality analyses indicate that in 10 out of 

28 the 12 urban case study areas a revised standard with a level of 65 ppb would be expected to 

29 maintain long-term O3 concentrations below those where the study by Jerrett et al. (2009) 

30 indicates the most confidence in the reported association with respiratory mortality. This is 

31 compared to 6 out of 12 areas for the current standard and 9 out of 12 for a standard with a level 

32 of 70 ppb (Table 4-3). 

 

34As discussed above, most of the study cities evaluated by Katsouyanni et al. (2009) for mortality would have met 

the current standard and a revised standard with a level of 70 ppb. 

35Though the majority of cities evaluated, based on the 30 ppb cut point analysis, would have met a standard with a 

level of 65 ppb over the entire study period (Table 4-2). 
4-54

1 Compared to the current standard and a revised standard with a level of 70 ppb, the 

2 second draft HREA estimates that a revised standard with a level of 65 ppb would generally 

3 reduce O3-associated mortality and morbidity risks based on estimates that use the full 

4 distributions of ambient O3 concentrations. When summed across urban case study areas, a 

5 standard with a level of 65 ppb is estimated to reduce total mortality associated with short-term 

6 O3 exposures by about 10%, compared to the current standard (Figure 4-13). Risks of respiratory 

7 mortality associated with long-term exposures were reduced by up to about 10%, compared to 

8 the current standard. As with a level of 70 ppb discussed above, reductions are estimated to be 

9 larger and more consistent across urban case study areas for the year with relatively higher 

10 ambient O3 concentrations. 

11 Similar to a standard with a level of 70 ppb, risk estimates indicate that a standard with a 

12 level of 65 ppb more effectively reduces mortality associated with the upper portions of 

13 distributions of ambient O3 concentrations. For days with area-wide concentrations at or above 

14 40 ppb, a standard level of 65 ppb is estimated to reduce O3-associated total mortality by about 

15 40%, when summed across cities. For days with area-wide concentrations at or above 60 ppb, a 

16 standard level of 65 ppb is estimated to reduce O3-associated total mortality by more than 80% 

17 (Figure 4-13). 

18 In summary, the rationale to support a level of 65 ppb would emphasize the potential 

19 public health significance of exposures of concern to 60 and 70 ppb O3, as well as 80 ppb, and of 

20 O3-attributable health effects that could occur across the range of ambient O3 concentrations.

21 Compared to a level of 70 ppb, a level of 65 ppb would place more emphasis on evidence from

22 controlled human exposure studies showing lung function decrements and pulmonary 

23 inflammation in healthy adults at 60 ppb O3. A standard with a level of 65 ppb would also place 

24 more emphasis on the potential for mortality and morbidity to be caused by the relatively low 

25 ambient O3 concentrations present in locations that would have met a standard with a level of 70 

26 ppb, and on the potential for further reductions in health risks, beyond those estimated for a level 

27 of 70 ppb. In addition, compared to a standard with a level of 70 ppb, a level of 65 ppb would 

28 place relatively less emphasis on the uncertainties associated with the evidence for O3-

29 attributable health effects at lower exposure and ambient concentrations, and less emphasis on 

30 uncertainties in the public health significance of O3 exposure and risk estimates for lower 

31 ambient concentrations. 

32 We next consider a standard with a level of 60 ppb. A level of 60 ppb is well-below the 

33 lowest O3 exposure concentration that has been reported to elicit a wide range of potentially 

34 adverse respiratory effects in healthy adults (i.e., 80 ppb), as noted above. A level of 60 ppb is 

35 also well-below the concentration where the combined occurrence of respiratory symptoms and 

36 lung function decrements was observed, a combination judged adverse by the ATS (discussed in 
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1 section 3.1.3). A level of 60 ppb corresponds to the lowest exposure concentration demonstrated 

2 to result in lung function decrements large enough to be judged an abnormal response by ATS, 

3 and that could be adverse in individuals with lung disease. A level of 60 ppb also corresponds to 

4 the lowest exposure concentration at which pulmonary inflammation has been reported. 

5 Based on the HREA analyses of O3 exposures of concern, a standard with a level of 60 

6 ppb is estimated to eliminate exposures of concern at or above the 70 and 80 ppb benchmark 

7 concentrations and to be more effective than higher standard levels at limiting exposures of 

8 concern at or above 60 ppb. On average over the years 2006 to 2010, a standard with a level of 

9 60 ppb is estimated to allow between 0 and about 1% of children in urban case study areas to 

10 experience exposures of concern at or above 60 ppb. This reflects a 96% reduction (on average 

11 across areas), compared to the current standard. A standard with a level of 60 ppb is estimated to 

12 allow virtually no children to experience two or more exposures of concern at or above 60 ppb. 

13 In the location and year with the highest exposure estimate, about 2% of children are estimated 

14 to experience exposures of concern at or above 60 ppb, with far less than 1% estimated to 

15 experience two or more such exposures (Table 4-4). 

16 Based on the HREA analyses of O3-induced lung function decrements, a standard with a 

17 level of 60 ppb would be expected to be more effective than a level of 70 or 65 ppb at limiting 

18 the occurrence of O3-induced lung function decrements. A standard with a level of 60 ppb is 

19 estimated to allow about 2% or less of children to experience one or more O3-induced lung 

20 function decrements ≥ 15% (almost 70% reduction, compared to current standard), and about 1% 

21 or less to experience two or more such decrements (3% in the location and year with the largest 

22 estimates). A standard set at 60 ppb is estimated to allow about 1% or less of children to 

23 experience large O3-induced lung function decrements (i.e., ≥ 20%), even in the worst-case 

24 locations and year (Table 4-5). 

25 On average over the years 2006 to 2010, a standard with a level of 60 ppb is estimated to 

26 allow about 5 to 11% of children, including asthmatic children, to experience one or more O3-

27 induced lung function decrements ≥ 10%. This reflects an average reduction of about 45%, 

28 compared to current standard. A standard with a level of 60 ppb is also estimated to allow about 

29 2 to 6% of children to experience two or more such decrements (51% reduction, compared to 

30 current standard). In the worst-case location and year, a standard set at 60 ppb is estimated to 

31 allow up to about 13% of children to experience one or more O3-induced lung function 

32 decrements ≥ 10%, and 7% to experience two or more such decrements (Table 4-5). 

33 A revised standard with a level of 60 ppb would be more effective than a level of 70 or 

34 65 ppb at maintaining short-term ambient O3 concentrations below those present in U.S. and 

35 Canadian epidemiologic study locations during study periods. Specifically, in all of the U.S. and 

36 Canadian epidemiologic studies evaluated, the majority of study cities had ambient O3
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1 concentrations that would likely have violated a standard with a level of 60 ppb. This includes 

2 the studies noted above, for which the majority of study cities would have met the current 

3 standard or a revised standard with a level of either 70 or 65 ppb (Tables 4-1 and 4-2). 

4 With regard to long-term O3 concentrations, air quality analyses indicate that in all of the 

5 urban case study areas evaluated a revised standard with a level of 60 ppb would be expected to 

6 maintain long-term O3 concentrations below those where the study by Jerrett et al. (2009) 

7 indicates the most confidence in the reported association with respiratory mortality. This is 

8 compared to 6 out of 12 areas for the current standard, 9 out of 12 for a standard with a level of 

9 70 ppb, and 10 out of 12 for a standard with a level of 65 ppb (Table 4-3). 

10 A standard with a level of 60 ppb would also be estimated to reduce epidemiology-based 

11 mortality and morbidity risks, compared to the current standard and potential alternative 

12 standards with levels of 70 or 65 ppb. Across urban case study areas, a standard set at 60 ppb is 

13 estimated to reduce total mortality associated with short-term O3 concentrations by up to 15%, 

14 and respiratory mortality associated with long-term O3 concentrations by up to 17%. Estimated 

15 risk reductions are larger for the model year with generally higher O3-associated risks (2007). 

16 When summed across urban case study areas, a standard set at 60 ppb is estimated to reduce total 

17 O3-associated mortality by about 10% (2009) and 14% (2007), compared to the current standard 

18 (Figure 4-13). 

19 As for standard levels of 70 and 65 ppb, risk estimates indicate that a standard with a 

20 level of 60 ppb more effectively reduces mortality associated with the upper portions of 

21 distributions of ambient O3 concentrations. For days with area-wide concentrations at or above 

22 40 ppb, a standard set at 60 ppb is estimated to reduce O3-associated total mortality by 

23 approximately 50%. For days with area-wide concentrations at or above 60 ppb, a standard level 

24 of 60 ppb is estimated to reduce O3-associated total mortality by over 95% (Figure 4-13). 

25 Overall we note that, compared to a standard with a level of 70 or 65 ppb, a level of 60 

26 ppb would place relatively more emphasis on the potential public health significance of the O3

27 exposures and O3-attributable health effects that could occur at lower ambient concentrations. A 

28 standard with a level of 60 ppb would reflect placing a relatively large emphasis on the evidence 

29 from controlled human exposure studies reporting lung function decrements and pulmonary 

30 inflammation in some healthy adults following exposures to 60 ppb O3, and relatively little 

31 emphasis on the uncertainties associated with the public health significance of these effects. A 

32 standard with a level of 60 ppb would also emphasize the importance of the small number of 

33 studies that reported health effect associations in locations that would have met a revised 

34 standard with a level of 65 ppb (Table 4-1), and on the potential for further reductions in health 

35 risks, beyond those estimated for a level of 65 ppb. In addition, compared to a standard with a 

36 level of 65 ppb, a level of 60 ppb would place relatively little emphasis on the uncertainties 
4-57

1 associated with the evidence for O3-attributable health effects at lower exposure and ambient 

2 concentrations, and little emphasis on uncertainties in the public health significance of O3

3 exposure and risk estimates for lower ambient concentrations. 

4 Based on our consideration of the available scientific evidence and exposure/risk 

5 information, we also conclude that standard levels below 60 ppb are not appropriate for 

6 consideration in this review. In reaching this conclusion, we take particular note of the following: 

7 The HREA estimates that meeting a standard with a level from 70 to 60 ppb would 

8 effectively reduce exposures of concern and lung function decrements. In particular, a 

9 level from the low end of this range could virtually eliminate exposures of concern, even 

10 for the lowest health benchmark concentration supported by the available controlled 

11 human exposure evidence (i.e., 60 ppb). To the extent lower exposure concentrations may 

12 result in adverse respiratory effects in some individuals, a standard level from 70 to 60 

13 ppb (particularly at or near 60 ppb) would be expected to also reduce exposures to O3

14 concentrations somewhat below 60 ppb. 

15

16 A revised standard with a level from 70 to 60 ppb would be more effective than the 

17 current standard at maintaining short- and long-term ambient O3 concentrations below 

18 those in locations that provided the basis for positive and statistically significant health 

19 effect associations in epidemiologic studies. In particular, in all of the U.S. and Canadian 

20 epidemiologic studies evaluated, the majority of study cities had ambient O3

21 concentrations that would likely have violated a standard with a level of 60 ppb. 

22

23 To the extent emphasis is placed on epidemiology-based mortality and/or morbidity risk 

24 estimates, meeting a standard with a level from 70 to 60 ppb would generally reduce such 

25 risks across the majority of the U.S. This is particularly the case for risks associated with 

26 ambient O3 concentrations from the upper portions of seasonal distributions. Given that 

27 uncertainty in quantitative estimates of O3-associated mortality and morbidity increases 

28 with the magnitude of NOX reductions required to simulate potential alternative 

29 standards, risk estimates associated with standard levels below 60 ppb would be 

30 associated with increasing uncertainty. 

31 Given the above, we conclude that, compared to standard levels from 70 to 60 ppb, the 

32 extent to which standard levels below 60 ppb could result in further public health improvements 

33 becomes notably less certain. Therefore, we conclude that it is not appropriate to consider 

34 standard levels below 60 ppb in this review. 

35 4.7 KEY UNCERTAINTIES AND AREAS FOR FUTURE RESEARCH AND 

36 DATA COLLECTION 

37 It is important to highlight the uncertainties associated with establishing standards for O3

38 during and after completion of the NAAQS review process. Research needs go beyond what is 

39 necessary to understand health effects, population exposures, and risks of exposure for purposes 
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1 of setting standards. Research can also support the development of more efficient and effective 

2 control strategies. In this section, we highlight areas for future health-related research, model 

3 development, and data collection activities to address these uncertainties and limitations in the 

4 current body of scientific evidence. 

5 As has been presented and discussed in the ISA, particularly chapters 4 through 7, the 

6 scientific body of evidence informing our understanding of health effects associated with long7 and short-term exposures to O3 has been broadened and strengthened since the O3 NAAQS

8 review completed in 2008. Still, we have concluded that O3 health research needs and priorities 

9 have not changed substantially since the 2007 O3 Staff Paper (EPA 2007). Key uncertainties and 

10 research needs that continue to be high priority for future reviews of the health-based standards 

11 are identified below: 

12 (1) An important aspect of risk characterization and decision making for air quality 

13 standard levels for the O3 NAAQS is the characterization of the shape of exposure-response 

14 functions for O3, including the identification of potential population threshold levels. Recent 

15 controlled human exposure studies of measurable lung function effects provide evidence for a 

16 smooth dose-response curve without evidence of a threshold for exposures between 40 and 120 

17 ppb O3 (US EPA, 2013, Figure 6-1). Considering the importance of estimating health risks in the 

18 range below 80 ppb O3, additional research is needed to evaluate responses in healthy and 

19 especially asthmatic individuals in the range of 40 to 70 ppb for 6-8 hr exposures while engaged 

20 in moderate exertion. 

21 (2) Similarly, for health endpoints reported in epidemiologic studies such as hospital 

22 admissions, ED visits, and premature mortality, an important aspect of characterizing risk is the 

23 shape of concentration-response functions for O3, including identification of potential population 

24 threshold levels. Most of the recent studies and analyses continue to show no evidence for a clear 

25 threshold in the relationships between O3 concentrations commonly observed in the U.S. during 

26 the O3 season and these health endpoints, though evidence indicates less certainty in the shape of 

27 the concentration-response curve at the lower end of the distribution of O3 concentrations. 

28 However, there continues to be heterogeneity in the O3-mortality relationship across cities (or 

29 regions), including effect modifiers that are also expected to vary regionally, which are sources 

30 of uncertainty. Additionally, whether or not exposure errors, misclassification of exposure, or 

31 potential impacts of other copollutants may be obscuring potential population thresholds is still 

32 unknown. 

33 (3) The extent to which the broad mix of photochemical oxidants and more generally 

34 other copollutants in the ambient air (e.g., PM, NO2, SO2, etc.) may play a role in modifying or 

35 contributing to the observed associations between ambient O3 and various morbidity effects and 

36 mortality continues to be an important research question. Ozone has long been known as an 
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1 indicator of health effects of the entire photochemical oxidant mix in the ambient air and has 

2 served as a surrogate for control purposes. A better understanding of sources of the broader 

3 pollutant mix, of human exposures, and of how other pollutants may modify or contribute to the 

4 health effects of O3 in the ambient air, and vice versa, is needed to better inform future NAAQS 

5 reviews. 

6 (4) As epidemiologic research has continued to be an important factor in assessing the 

7 public health impacts of O3, methodological issues in epidemiologic studies have received 

8 greater visibility and scrutiny. There remains a need to further examine alternative modeling 

9 specifications and control of time-varying factors, and to better understand the role of 

10 copollutants in the ambient air. Additionally, there remains uncertainty around the role of 

11 temperature as a potential confounder or effect modifier in epidemiologic models. 

12 (5) Recent animal toxicological evidence, combined with limited evidence from 

13 controlled human exposure studies of cardiovascular morbidity and epidemiologic studies of 

14 cardiovascular mortality, have provided evidence of both direct and indirect effects on the 

15 cardiovascular system. However, additional work will need to examine biologically plausible 

16 mechanisms of cardiovascular effects, expand upon preliminary evidence from controlled human 

17 exposure studies, address inconsistencies observed in epidemiologic studies of cardiovascular 

18 morbidity, and determine the extent to which O3 is directly implicated or works together with 

19 other pollutants in causing adverse cardiovascular effects in both at-risk and the general 

20 populations. 

21 (6) Most epidemiologic studies of short-term exposure effects have employed time-series 

22 or case-crossover study designs and have been conducted in large populations. These study 

23 designs remain subject to uncertainty due to use of ambient fixed-site data serving as a surrogate 

24 for ambient exposures, and to the difficulty of determining the impact of any single pollutant 

25 among the mix of pollutants in the ambient air. Measurements made at stationary outdoor 

26 monitors have been used as independent variables for air pollution, but the accuracy with which 

27 these measurements actually reflect subjects’ exposure is not yet fully understood. Also, 

28 additional research is needed to improve the characterization of the degree to which discrepancy 

29 between stationary monitor measurements and actual pollutant exposures introduces error into 

30 statistical estimates of pollutant effects in epidemiologic studies. 

31 (7) Recent studies of “long-term” O3 often evaluate associations with daily maximum 

32 concentrations, averaged over the O3 season. Research is needed to better understand the extent 

33 to which health effects associated with such long-term metrics are attributable to long-term 

34 average concentrations versus the repeated occurrence of daily maximum concentrations. 

35 (8) Improved understanding of human exposures to ambient O3 and to related 

36 copollutants is an important research need. Population-based information on human exposure for 
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1 healthy adults and children and at-risk populations, including people with asthma, to ambient O3

2 concentrations, including exposure information in various microenvironments, is needed to better 

3 evaluate current and future O3 exposure models. Such information is needed for sufficient 

4 periods to facilitate evaluation of exposure models throughout the O3 season. 

5 (9) Information is needed to improve inputs to current and future population-based O3

6 exposure and health risk assessment models. Collection of time-activity data over longer time 

7 periods is needed to reduce uncertainty in the modeled exposure distributions that form an 

8 important part of the basis for decisions regarding NAAQS for O3 and other air pollutants. 

9 Research addressing energy expenditure and associated breathing rates in various population 

10 groups, particularly healthy children and children with asthma, in various locations, across the 

11 spectrum of physical activity, including sleep to vigorous exertion, is needed. 

12 (10) An important consideration in the O3 NAAQS review is the characterization of 

13 background levels. There still remain substantial uncertainties in the characterization of 8-hr 

14 daily max O3 background concentrations. Further research to improve the evaluation of the 

15 global and regional models which have been used to characterize estimates of background levels 

16 would improve understanding of the role of non-U.S. anthropogenic emissions on O3 levels over 

17 the U.S. 

18 4.8 SUMMARY OF PRELIMINARY STAFF CONCLUSIONS ON PRIMARY 

19 STANDARD 

20 In this section, we summarize our preliminary conclusions regarding the primary O3

21 standard. Staff conclusions in the final PA will be informed by our consideration of the available 

22 scientific evidence as assessed in the ISA, air quality/exposure/risk information assessed in the 

23 final HREA, recommendations received from CASAC based on their review of this second draft 

24 document, and comments received from members of the public. 

25 As an initial matter, as discussed in section 3.4, staff concludes that reducing precursor 

26 emissions to achieve O3 concentrations that meet the current standard will provide important 

27 improvements in public health protection. This initial conclusion is based on (1) the strong body 

28 of scientific evidence indicating a wide range of adverse health outcomes attributable to 

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

30 O3 exposures and health risks upon meeting the current standard, compared to recent air quality. 

31 Strong support for this conclusion is provided by the available health evidence and HREA 

32 estimates of O3 exposures of concern and O3-induced lung function risks. Some support for this 

33 conclusion is also provided by HREA estimates of O3-associated mortality and morbidity. 

34 In considering the available evidence and information, staff further concludes that the O3-

35 attributable health effects estimated to be allowed by air quality that meets the current primary 
4-61

1 standard for O3 can reasonably be judged important from a public health perspective. Thus, we 

2 conclude that the available health evidence and exposure/risk information call into question the 

3 adequacy of the public health protection provided by the current standard. This conclusion is 

4 based on consideration of the scientific evidence assessed in the ISA, including controlled human 

5 exposure and epidemiologic studies, as well as animal toxicology studies; the air quality, 

6 exposure, and risk analyses presented in the second draft HREA for air quality that just meets the 

7 current standard; and advice received from CASAC in their review of the first draft PA and in 

8 previous reviews. 

9 In reaching the above conclusion regarding the current standard, we also reach 

10 preliminary conclusions for the Administrator’s consideration in making decisions on the 

11 elements of a potential alternative primary O3 standard, as summarized below. We recognize that 

12 selecting from among potential alternative standards will necessarily reflect consideration of 

13 qualitative and quantitative uncertainties inherent in the relevant evidence and in the assumptions 

14 of the quantitative exposure and risk assessments. Any such standard should protect public health 

15 against health effects associated with exposure to O3, alone or in combination with related 

16 photochemical oxidants, taking into account both evidence-based and exposure- and risk-based 

17 considerations, and the nature and degree of uncertainties in such information. In reaching 

18 conclusions about these ranges of potential alternative standards for consideration, we are 

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

20 requisite to protect public health with an adequate margin of safety. The primary standards are to 

21 be neither more nor less stringent than necessary. Thus, the Act does not require that primary 

22 NAAQS be set at zero-risk levels, but rather at levels that reduce risk sufficiently to protect 

23 public health with an adequate margin of safety. 

24 The degree of public health protection provided by the standard is due to the collective 

25 impact of the elements of the standard, including the indicator, averaging time, level, and form. 

26 Staff’s preliminary conclusions on each of these elements are summarized below. 

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

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

29 photochemical oxidants. Based on the available information staff preliminarily concludes 

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

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

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

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

34 alone. Staff notes that control of ambient O3 levels is generally understood to provide the 

35 best means of controlling photochemical oxidants in general, and thus of protecting 
4-62

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

2 photochemical oxidants, independent of effects specifically related to O3.

3

4 (2) It is appropriate to continue to use an 8-hour averaging time for the primary O3 standard. 

5 (a) Staff preliminarily concludes that an 8-hour averaging time remains appropriate 

6 for addressing health effects associated with short-term exposures to ambient O3. 

7 An 8-hour averaging time is similar to the exposure periods evaluated in 

8 controlled human exposure studies, including recent studies that provide evidence 

9 for respiratory effects following exposures to O3 concentrations below the level of 

10 the current standard. In addition, epidemiologic studies provide evidence for 

11 health effect associations with 8-hour O3 concentrations, as well as with 1-hour 

12 and 24-hour concentrations. A standard with an 8-hour averaging time (combined 

13 with an appropriate standard form and level) would also be expected to provide 

14 substantial protection against health effects attributable to 1- and 24-hour 

15 exposures. 

16

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

18 can provide protection against respiratory effects associated with longer term O3

19 exposures. Analyses in the HREA show that as air quality is adjusted to just meet 

20 the current 8-hour standard, most study areas are estimated to experience 

21 reductions in respiratory mortality associated with long-term O3 concentrations, 

22 indicating that an O3 standard with an 8-hour averaging time can reduce 

23 respiratory mortality reported to be associated with long-term O3 concentrations. 

24 Moreover, the large majority of the U.S. population lives in locations where 

25 reducing NOX emissions would be expected to result in modest decreases in warm 

26 season averages of daily 8-hour ambient O3 concentrations. This suggests that 

27 reductions in precursor emissions in order to meet a standard with an 8-hour 

28 averaging time would also be expected to reduce the types of long-term O3

29 concentrations that have been associated with respiratory morbidity in 

30 epidemiologic studies. 

31

32 In addition, an analysis in the PA of whether just meeting the current or 

33 alternative O3 standards, with 8-hour averaging times, would be expected to 

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

35 below those present in most of the cities that provided the basis for a positive and 

36 statistically significant association with respiratory mortality (Jerrett et al., 2009). 
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1 This suggests that a standard with an 8-hour averaging time can maintain seasonal 

2 averages of 1-hour daily max O3 concentrations below those that provided the 

3 basis for the association with respiratory mortality (and below the concentration at 

4 which study authors noted limited evidence of an effects threshold). Taken 

5 together, these analyses suggest that a standard with an 8-hour averaging time, 

coupled with the current 4th 6 -highest form and an appropriate level, could provide 

7 appropriate protection against the long-term O3 concentrations reported to be 

8 associated with respiratory morbidity and mortality. 

9

10 (3) It is appropriate to revise the level of the standard since the evidence and information 

11 from the exposure and risk assessments in this review provide strong support for 

12 consideration of an O3 standard with a level that would provide increased health 

13 protection for at-risk groups, including people with asthma, especially children; the 

14 lifestages of children and older adults; people with certain genetic variants; people with 

15 reduced intake of certain nutrients; and outdoor workers against an array of adverse 

16 health effects. These health effects range from decreased lung function, pulmonary 

17 inflammation, and respiratory symptoms to serious indicators of respiratory morbidity 

18 including ED visits and hospital admissions for respiratory causes, respiratory and all19 cause, non-accidental mortality. We also conclude that exposures of concern and health 

20 risks projected to remain upon meeting the current standard, based on the exposure and 

21 risk assessments, are indicative of risks to these populations and lifestages that can 

22 reasonably be judged to be important from a public health perspective. This reinforces 

23 our conclusion that consideration should be given to revising the level of the standard so 

24 as to provide increased public health protection. 

25

26 It is appropriate to consider a standard level within the range of 70 ppb to 60 ppb, 

27 reflecting our judgment that a standard set within this range could provide an appropriate 

28 degree of public health protection and would result in important improvements in 

29 protecting the health of at-risk populations and lifestages. Standard levels within this 

30 range that were considered in staff analyses of air quality, exposure, and risk include 70, 

31 65 and 60 ppb, representative of levels within the upper, middle, and lower parts of this 

32 range, respectively. Further, it would not be appropriate to consider increasing the level 

33 of the current standard, thereby decreasing public health protection. 

34

(4) It is appropriate to continue to use the 4th 35 -highest daily max form of the standard. Staff 

notes that the 4th 36 -highest daily max was selected in 1997 in recognition of the public 
4-64

1 health protection provided by this form, when coupled with an appropriate averaging 

2 time and level, combined with its stability for implementation programs. The currently 

3 available evidence and exposure/risk information does not call into question these 

4 conclusions from previous reviews. Therefore, we reach the preliminary conclusion that 

it is appropriate to consider retaining the current 4th 5 -highest daily max form for an O3

6 standard with an 8-hour averaging time and a revised level, as discussed above. 

7 4.9 REFERENCES 

8 Adams, W. C. (2006) Comparison of chamber 6.6 hour exposures to 0.04-0.08 ppm ozone via square-wave and 

9 triangular profiles on pulmonary responses. Inhalation Toxicol. 18: 127-136. 

10 http://dx.doi.org/10.1080/08958370500306107 

11 Bell, ML; Peng, RD; Dominici, F. (2006). The exposure-response curve for ozone and risk of mortality and the 

12 adequacy of current ozone regulations. Environ Health Perspect 114: 532-536.

13 Brown, JS; Bateson, TF; McDonnell, WF. (2008). Effects of exposure to 0.06 ppm ozone on FEV1 in humans: A 

14 secondary analysis of existing data. Environ Health Perspect 116: 1023-1026. 

15 http://dx.doi.org/10.1289/ehp.11396 

16 Cakmak, S; Dales, RE; Judek, S. (2006). Respiratory health effects of air pollution gases: Modification by education 

17 and income. Arch Environ Occup Health 61: 5-10. http://dx.doi.org/10.3200/AEOH.61.1.5-10

18 Dales, RE; Cakmak, S; Doiron, MS. (2006). Gaseous air pollutants and hospitalization for respiratory disease in the 

19 neonatal period. Environ Health Perspect 114: 1751-1754. http://dx.doi.org/10.1289/ehp.9044 

20 Darrow, LA; Klein, M; Sarnat, JA; Mulholland, JA; Strickland, MJ; Sarnat, SE; Russell, AG; Tolbert, PE. (2011a). 

21 The use of alternative pollutant metrics in time-series studies of ambient air pollution and respiratory 

22 emergency department visits. J Expo Sci Environ Epidemiol 21: 10-19. 

23 http://dx.doi.org/10.1038/jes.2009.49 

24 Frey, C.; Samet, J.M. (2012). CASAC Review of the EPA’s Policy Assessment for the Review of the Ozone National 

25 Ambient Air Quality Standards (First External Review Draft – August 2012). EPA-CASAC-13-003. 

26 November 26, 2012. Available online at: 

27 Jerrett, M; Burnett, RT; Pope, CA, III; Ito, K; Thurston, G; Krewski, D; Shi, Y; Calle, E; Thun, M. (2009). Long28 term ozone exposure and mortality. N Engl J Med 360: 1085-1095. 

29 http://dx.doi.org/10.1056/NEJMoa0803894 

30 Katsouyanni, K; Samet, JM; Anderson, HR; Atkinson, R; Le Tertre, A; Medina, S; Samoli, E; Touloumi, G; 

31 Burnett, RT; Krewski, D; Ramsay, T; Dominici, F; Peng, RD; Schwartz, J; Zanobetti, A. (2009). Air 

32 pollution and health: A European and North American approach (APHENA). (Research Report 142). 

33 Boston, MA: Health Effects Institute. http://pubs.healtheffects.org/view.php?id=327 

34 Kim, CS; Alexis, NE; Rappold, AG; Kehrl, H; Hazucha, MJ; Lay, JC; Schmitt, MT; Case, M; Devlin, RB; Peden, 

35 DB; Diaz-Sanchez, D. (2011). Lung function and inflammatory responses in healthy young adults exposed 

36 to 0.06 ppm ozone for 6.6 hours. Am J Respir Crit Care Med 183: 1215-1221. 

37 http://dx.doi.org/10.1164/rccm.201011-1813OC 

38 Mar, TF; Koenig, JQ. (2009). Relationship between visits to emergency departments for asthma and ozone exposure 

39 in greater Seattle, Washington. Ann Allergy Asthma Immunol 103: 474-479. 
4-65

1 McDonnell, WF; Stewart, PW; Smith, MV. (2010). Prediction of ozone-induced lung function responses in humans. 

2 Inhal Toxicol 22: 160-168. http://dx.doi.org/10.3109/08958370903089557 

3 Samet, J.M. (2010) Review of EPA’s Proposed Ozone National Ambient Air Quality Standard (Federal Register, 

4 Vol. 75, Nov. 11, January 19, 2010). EPA-CASAC-10-007. February 19, 2010. Available online at: 

5 Samet, J.M. (2011) Clean Air Scientific Advisory Committee (CASAC) Response to Charge Questions on the 

6 Reconsideration of the 2008 Ozone National Ambient Air Quality Standards. EPA-CASAC-11-004. 

7 March 30, 2011. Available online at: 

8 http://yosemite.epa.gov/sab/sabproduct.nsf/0/F08BEB48C1139E2A8525785E006909AC/$File/EPA9 CASAC-11-004-unsigned+.pdf 

10 Schelegle, ES; Morales, CA; Walby, WF; Marion, S; Allen, RP. (2009). 6.6-hour inhalation of ozone concentrations 

11 from 60 to 87 parts per billion in healthy humans. Am J Respir Crit Care Med 180: 265-272. 

12 http://dx.doi.org/10.1164/rccm.200809-1484OC 

13 Smith, RL; Xu, B; Switzer, P. (2009b). Reassessing the relationship between ozone and short-term mortality in U.S. 

14 urban communities. Inhal Toxicol 21: 37-61. http://dx.doi.org/10.1080/08958370903161612 

15 Stieb, DM; Szyszkowicz, M; Rowe, BH; Leech, JA. (2009). Air pollution and emergency department visits for 

16 cardiac and respiratory conditions: A multi-city time-series analysis. Environ Health Global Access Sci 

17 Source 8: 25. http://dx.doi.org/10.1186/1476-069X-8-25 

18 U.S. Environmental Protection Agency. (2012). Health Risk and Exposure Assessment for Ozone, First External 

19 Review Draft, U.S. Environmental Protection Agency, Research Triangle Park, NC. EPA 452/P-12-001. 

20 U.S. Environmental Protection Agency. (2013). Integrated Science Assessment for Ozone and Related 

21 Photochemical Oxidants (Final Report). U.S. Environmental Protection Agency, Washington, DC, 

22 EPA/600/R-10/076F. 

23 U.S. Environmental Protection Agency. (2014). Health Risk and Exposure Assessment for Ozone, Second External 

24 Review Draft. Office of Air Quality Planning and Standards, Research Triangle Park, NC. EPA-452/P-14-

25 004a. 
5-1 

1 5 ADEQUACY OF THE CURRENT SECONDARY STANDARD

2 This chapter presents staff’s considerations and preliminary conclusions regarding the 

3 adequacy of the current secondary O3 NAAQS. In doing so, we pose the following overarching 

4 question: 

5 • Does the currently available scientific evidence- and exposure/risk-based 

6 information, as reflected in the ISA and WREA, support or call into question the 

7 adequacy and/or appropriateness of the protection afforded by the current 

8 secondary O3 standard? 

9 In addressing this overarching question, we pose a series of more specific questions, as 

10 discussed in sections 5.1 through 5.5 below. We consider the nature of O3-induced effects, 

11 including the nature of the exposures that drive the biological and ecological response and 

12 related biologically-relevant exposure metrics (section 5.1); the scientific evidence and 

13 exposure/risk information, including that of associated ecosystem services, regarding (a) tree 

14 growth, productivity and carbon storage (section 5.2), (b) crop yield loss (section 5.3), (c) visible 

15 foliar injury (section 5.4), and (d) other welfare effects (section 5.5). Section 5.6 describes 

16 advice and recommendations received from CASAC. In section 5.7, we revisit the overarching 

17 question of this chapter and present preliminary staff conclusions on the adequacy of the current 

18 secondary standard. 

19 5.1 NATURE OF EFFECTS AND BIOLOGICALLY-RELEVANT EXPOSURE

20 METRIC 

21 • Does the current evidence alter our conclusions from the previous review

22 regarding the nature of O3-induced welfare effects?

23 The current body of O3 welfare effects evidence confirms the conclusions reached in the 

24 last review on the nature of O3-induced welfare effects and is summarized in the ISA as follows 

25 (U.S. EPA, 2013, p. 1-8):

26 The welfare effects of O3 can be observed across spatial scales, starting at the 

27 subcellular and cellular level, then the whole plant and finally, ecosystem-level 

28 processes. Ozone effects at small spatial scales, such as the leaf of an individual 

29 plant, can result in effects along a continuum of larger spatial scales. These 

30 effects include altered rates of leaf gas exchange, growth, and reproduction at the 

31 individual plant level, and can result in broad changes in ecosystems, such as 

32 productivity, carbon storage, water cycling, nutrient cycling, and community 

33 composition. 
5-2 

1 This body of evidence has been expanding since the phytotoxic effects of O3 were first 

2 identified by Richards, et al in 1958 who showed that “O3 was a constituent of smog that caused 

3 foliar injury on grapes in California” (Flagler, 1998). In the half century that has followed, a 

4 large number of studies have been conducted both in and outside of the U.S. to examine the 

5 impacts of O3 on plants and their associated ecosystems. Taken together, these studies 

6 demonstrate that O3 -induced effects that occur at the subcellular and cellular levels, at sufficient 

7 magnitudes propagate up to produce larger scale effects that affect the whole organism. In 

8 addition, the studies assessed in this review have further increased our understanding of the 

9 molecular, biochemical and physiological mechanisms that explain how plants are affected by 

10 O3, in the absence of other stressors, particularly in the area of genomics (U.S. EPA, 2013, 

11 Chapter 9, section 9.3). Based on its assessment of this extensive body of science, the ISA 

12 determined that a causal relationship exists between exposure to O3 in ambient air and visible 

13 foliar injury effects on vegetation, reduced vegetation growth, reduced productivity in terrestrial 

14 ecosystems, reduced yield and quality of agricultural crops and alteration of below-ground 

15 biogeochemical cycles (U.S. EPA 2013, Table 1-2). Additionally, the ISA determined that a 

16 likely to be causal relationship exists between exposures to O3 in ambient air and reduced carbon 

17 sequestration in terrestrial ecosystems, alteration of terrestrial ecosystem water cycling and 

18 alteration of terrestrial community composition (U.S. EPA, 2013, Table 1-2). 

19 From this set of effects that the ISA has concluded to be causally or likely causally

20 related to O3 in ambient air, we focus primarily on three categories of effects (i.e., visible foliar 

21 injury; impacts on tree growth, productivity and carbon storage; and crop yield loss). As 

22 recognized in the ISA, controlled exposure studies, which are the best method for isolating or 

23 characterizing the role of O3 in inducing the observed plant effects, “have clearly shown that 

24 exposure to O3 is causally linked to visible foliar injury, decreased photosynthesis, changes in 

25 reproduction, and decreased growth” in vegetative species (U.S. EPA 2013, p. 1-15). These plant 

26 level effects are also linked to a cascade of other ecosystem level effects. For example, studies at 

27 larger spatial scales support the controlled exposure study results and indicate that “ambient O3

28 exposures can effect ecosystem productivity, crop yield, water cycling, and ecosystem 

29 community composition” (U.S. EPA 2013, p. 1-15). Thus, O3 effects on vegetation may have 

30 implications at the individual, species, population and whole ecosystem level, including 

associated ecosystem services. 1 31 

 

1 Ecosystem services are the benefits that people obtain from ecosystems and have been stated to include 

“provisioning services such as food and water; regulating services such as regulation of floods, drought, land 

degradation, and disease; supporting services such as soil formation and nutrient cycling; and cultural services such 

as recreational, spiritual, religious, and other nonmaterial benefits” according to the Millennium Ecosystem 

Assessment (UNEP, 2003; U.S. EPA, 2013, p. lxxii).
5-3 

1 Our consideration of O3 welfare effects and the significance or weight to place on a given 

2 study or exposure/risk assessment result, is informed by the understanding, based on the entire 

3 body of vegetation effects science, that a variety of other factors can either mitigate or exacerbate 

4 the O3-plant interactions and are recognized sources of uncertainty and variability. These 

5 include: 1) multiple genetically-influenced determinants of O3 sensitivity; 2) changing sensitivity 

6 across growth stages; 3) co-occurring stressors and/or modifying environmental factors; 4) a 

7 paucity of information on most of the 43,000 U.S. plant species (U.S. EPA 2013, section 9.4.8; 

8 U.S. EPA 2006; U.S. EPA, 2007, section 7.4.2). 

9 • Does the current evidence continue to support a cumulative, seasonal exposure 

10 index as a biological-relevant and appropriate metric for assessment of the 

11 evidence and exposure/risk information? 

12 In this review, the ISA assessment of the full body of currently available evidence stated 

13 the following regarding biological indices (U.S. EPA, 2013, p. 2-44): 

14 The main conclusions from the 1996 and 2006 O3 AQCDs regarding indices 

15 based on ambient exposure remain valid. These key conclusions can be restated 

16 as follows: 

17 • ozone effects in plants are cumulative; 

18 • higher O3 concentrations appear to be more important than lower 

19 concentrations in eliciting a response; 

20 • plant sensitivity to O3 varies with time of day and plant development 

21 stage; 

22 • quantifying exposure with indices that cumulate hourly O3 concentrations 

23 and preferentially weight the higher concentrations improves the 

24 explanatory power of exposure/response models for growth and yield, 

25 over using indices based on mean and peak exposure values. 

26 Thus, the current evidence, as in other recent reviews, continues to support a cumulative, 

27 seasonal exposure index as a biologically-relevant and appropriate metric for assessment of the 

28 evidence and exposure/risk information. To this point, the available body of evidence provides a 

29 wealth of information on the aspects of O3 exposure that are most important in influencing plant 

30 response. While a variety of “factors with known or suspected bearing on the exposure-response 

31 relationship, including concentration, time of day, respite time, frequency of peak occurrence, 

32 plant phenology, predisposition, etc.,” have been identified (U.S. EPA, 2013, section 9.5.2), the 

33 importance of the duration of the exposure and the relatively greater importance of higher 

34 concentrations over lower in determining plant response to O3 have been well documented (U.S. 

35 EPA, 2013, section 9.5.3). Much of this work was completed by the mid-1990s, and was 
5-4 

1 summarized in the 1996 Criteria Document (U.S. EPA, 1996), while the additional newer work 

2 is assessed in the subsequent 2006 Criteria Document and 2013 ISA. 

3 In conjunction with this research on plant response to O3 exposures, others have 

4 developed “mathematical approaches for summarizing ambient air quality information in 

5 biologically meaningful forms for O3 vegetation effects assessment purposes ...” (U.S. EPA, 

6 2013, section 9.5.3). A large set of exposure indices have been developed that use a variety of 

7 functions to weight factors that have been shown to influence vegetation exposure-response 

8 relationships (U.S. EPA, 2013, section 9.5.2). As discussed in the ISA, several indices have 

9 been developed to attempt to incorporate some of the biological, environmental, and exposure 

10 factors that influence the magnitude of the biological response and contribute to observed 

11 variability (U.S. EPA, 2013, section 9.5.2). As with any summary statistic, these exposure 

12 indices retain information on some, but not all, characteristics of the original observations. 

13 Flux models have been identified in recent years to take into account more of the factors 

14 that influence the response and contribute to observed variability (U.S. EPA, 2013, section 

15 9.5.4). We note that “some researchers have claimed that using flux models can be used to better 

16 predict vegetation responses to O3 than exposure-based approaches...” because flux models 

17 estimate the ambient O3 concentration that actually enters the leaf (i.e., flux or deposition) (U.S. 

18 EPA, 2013, p. 9-114). However, “[f]lux calculations are data intensive and must be carefully 

19 implemented” (U.S. EPA, 2013, p. 9-114). Further, “[t]his uptake-based approach to quantify 

20 the vegetation impact of O3 requires inclusion of those factors that control the diurnal and 

21 seasonal O3 flux to vegetation (e.g., climate patterns, species and/or vegetation-type factors and 

22 site-specific factors)” (U.S. EPA, 2013, p. 9-114). Each species has different amounts of internal 

23 detoxification potential that may protect species to differing degrees. This balance between O3

24 flux and detoxification processes has been termed the “effective flux”. Accordingly, the 

25 “models have to distinguish between stomatal and non-stomatal components of O3 deposition to 

26 adequately estimate actual concentration reaching the target tissue of a plant to elicit a response” 

27 and “[d]etermining this O3 uptake via canopy and stomatal conductance relies on models to 

28 predict flux and ultimately the ‘effective’ flux” (U.S. EPA, 2013, pp. 9-114). The lack of 

29 detailed species- and site-specific data required for flux modeling in the U.S. and the lack of 

30 understanding of detoxification processes have made this technique less viable for use in 

31 vulnerability and risk assessments at the national scale in the U.S. (U.S. EPA, 2013, section 

32 9.5.4). 

33 Thus, in the last two reviews of the O3 secondary standard, completed in 1997 and 2008, 

34 the EPA concluded that the risk to vegetation comes primarily from cumulative exposures to O3

35 over a season or seasons and, in both reviews, the EPA proposed, as one alternative, a secondary 

36 standard set in terms of such a form (61 FR 65716, 72 FR 37818). Although in both reviews the 
5-5 

1 secondary standard was revised to be identical to a revised primary standard (with an 8-hour 

2 averaging time), the Administrator, in both cases, also concluded, consistent with CASAC 

3 advice, that a cumulative, seasonal index was the most biologically relevant way to relate 

4 exposure to plant growth response (62 FR 38856, 73 FR 16436). Most recently, in the 2010 

5 proposed reconsideration of the 2008 decision, the EPA again proposed to conclude that O3

6 exposure indices that cumulate differentially weighted hourly concentrations are the best 

7 candidates for relating exposure to plant growth responses and proposed to set the secondary 

8 standard only in terms of one such form, the W126 (75 FR 2938). 

9 Based on the long-established conclusions and long-standing supporting evidence 

10 described above, we continue to focus on the aspects of ambient O3 exposures that have 

11 biological relevance and the biologically-relevant exposure indices or metrics that have been 

12 designed in light of this consideration, i.e., cumulative seasonal indices. Since the review 

13 completed in 1997, which was the first to focus on cumulative indices, attention has been given 

14 primarily to two different cumulative index forms: SUM06 and W126. The SUM06 index is a 

15 threshold-based approach described as the sum of all hourly O3 concentrations greater or equal to 

16 0.06 ppm observed during a specified daily and seasonal time window (U.S. EPA, 2013, section 

17 9.5.2). The W126 index is a non-threshold approach described as the sigmoidally weighted sum 

18 of all hourly O3 concentrations observed during a specified daily and seasonal time window, 

19 where each hourly O3 concentration is given a weight that increases from 0 to 1 with increasing 

20 concentration (Lefohn et al, 1988; Lefohn and Runeckles, 1987; U.S. EPA, 2013, section 9.5.2). 

21 The EPA used the W126 index to consider welfare effects in the last review, as well as the 2010 

22 proposed reconsideration of the 2008 decision; this approach received support from CASAC, as 

23 discussed in section 5.6 below. Consistent with the ISA conclusions regarding the 

24 appropriateness of considering cumulative exposure indices for O3 effects of concern based on 

25 the evidence available in this review, we again conclude that the current evidence continues to 

26 support a cumulative, seasonal exposure index as a biologically-relevant and appropriate metric 

27 for assessment of the evidence and exposure/risk information, and in particular, the W126 

28 cumulative, seasonal metric (U.S. EPA, 2013, p. 2-44 and section 9.5.2). 

29 • Within what paradigm may it be appropriate to consider the potential adversity 

30 of public welfare effects of O3? 

31 The Clean Air Act requires that a secondary standard be protective against those known 

32 or anticipated O3 effects that are “adverse” to the public welfare, not all identifiable O3-induced 

33 effects. Unlike the use of the terms adverse, injury or damage in the scientific literature, in the 

34 NAAQS policy context, these terms have been interpreted in a particular way. Specifically, O3-

35 induced “injury” to vegetation has been defined as encompassing all plant reactions, including 
5-6 

1 reversible changes or changes in plant metabolism (e.g., altered photosynthetic rate), altered 

2 plant quality or reduced growth that does not impair the intended use or value of the plant. In 

3 contrast, “damage” has been defined to include only those injury effects that reach sufficient 

4 magnitude as to also reduce or impair the intended use or value of the plant to the public and thus 

5 potentially become adverse to the public welfare. Examples of vegetation effects that have been 

6 classified as damage include reductions in aesthetic values (e.g., foliar injury in ornamental 

7 species) as well as losses in terms of weight, number, or size of harvestable plant parts. Biomass 

8 loss in tree species can also be considered damage or adverse to the public welfare if it includes 

9 slower growth in species harvested for timber or other fiber uses. In the context of evaluating 

10 effects on single plants or species grown in monocultures such as managed forests, this construct 

11 continues to remain useful (73 FR 16492/96). 

12 However, given the increasing scientific literature linking O3 effects on plants or species 

13 to effects at the community or ecosystem level, a more expansive construct or paradigm of what 

14 constitutes O3 “damage” beyond that of the individual or species level is appropriate. A number 

15 of broader paradigms have been discussed in the literature (72 FR 37890; Hogsett et al., 1997; 

16 Young and Sanzone, 2002). In the 2008 review, the Administrator expressed support for relying 

17 on a definition of “adverse” discussed in section IV.A.3 of the proposal (62 FR 37889-37890) 

18 that embeds “the concept of ‘intended use’ of the ecological receptors and resources that are 

19 affected, and applies that concept beyond the species level to the ecosystem level” (73 FR 

20 16496). For example, in the 2008 rulemaking notice, the Administrator took note of “a number 

21 of actions taken by Congress to establish public lands that are set aside for specific uses that are 

22 intended to provide benefits to the public welfare, including lands that are to be protected so as to 

23 conserve the scenic value and the natural vegetation and wildlife within such areas, and to leave 

24 them unimpaired for the enjoyment of future generations” (73 FR 16496). 

25 Since the 2008 O3 review, our approach to assessing adversity to the public welfare in 

26 NAAQS reviews has continued to evolve. In particular, we consider the concept of ecosystem 

27 services in a broader paradigm. An extensive look at the range of services than can be provided 

28 by ecosystems is described in the Millennium Ecosystem Assessment (MEA, 2005). Ecosystem 

29 services can be generally defined as the benefits that individuals and organizations obtain from 

30 ecosystems. The EPA has previously defined ecological goods and services for the purposes of 

31 benefits assessment as the ‘‘outputs of ecological functions or processes that directly or 

32 indirectly contribute to social welfare or have the potential to do so in the future. Some outputs 

33 may be bought and sold, but most are not marketed’’ and has relied on this definition in 

34 regulatory impact analyses for previous NAAQS reviews (U.S. EPA, 2006b). In the review of 

35 the secondary NAAQS for oxides of nitrogen and sulfur, EPA recognized that changes in 

36 ecosystem services may be used to aid in characterizing a known or anticipated adverse effect to 
5-7 

1 public welfare and that an evaluation of adversity to the public welfare might consider the 

2 likelihood, type, magnitude, and spatial scale of the effect, as well as the potential for recovery 

3 and any uncertainties relating to these conditions (77 FR 20232). In the context of this review, 

4 ecosystem services are being evaluated and assessed in the REA as one way to characterize the 

5 possible public welfare benefits received from ecosystem resources and how those services 

6 might be expected to change under air quality scenarios representing the current and potential 

7 alternative secondary standards (U.S. EPA, 2014, chapter 5). Thus, in considering the evidence

8 and exposure risk information associated with the welfare endpoints identified below, in the 

9 context of consideration of adequacy of the current standard in this chapter and potential 

10 alternative standards in chapter 6, we consider also how they fit within this paradigm. 

11 5.2 FOREST TREE GROWTH, PRODUCTIVITY AND CARBON STORAGE

12 This section considers the current evidence and exposure/risk information to inform 

13 consideration of the adequacy of the protection provided by the current standard from known and 

14 anticipated adverse welfare effects of O3 related to growth, productivity, and carbon storage of 

15 forest trees, and other associated effects. Trees are important from a public welfare perspective 

16 because they provide many valued services to humans. In addition to the aesthetic value 

17 discussed below in 5.4, these include: food, fiber, timber, other forest products, habitat, 

18 recreational opportunities, climate regulation, erosion control, air pollution removal, hydrologic 

19 and fire regime stabilization (U.S. EPA, 2014, section 6.1, Figure 6-1, section 6.4, Table 6-12). 

20 This section includes a discussion of the policy-relevant evidence and weight-of-evidence 

21 conclusions discussed in the ISA (section 5.2.1) and the exposure/risk results, including 

22 associated ecosystem services (section 5.2.2) described in the second draft WREA. Important 

23 uncertainties and limitations in the available information are discussed throughout the sections. 

24 These discussions highlight the information we consider relevant to answering the overarching 

25 question and associated policy-relevant questions included in this section. 

26 5.2.1 Evidence-based Considerations 

27 • To what extent has scientific information become available that alters or 

28 substantiates our prior conclusions of O3-related effects on forest tree growth, 

29 productivity and carbon storage and of factors that influence associations between 

30 O3 concentrations and these effects?

31 Research published since the 2006 AQCD substantiates prior conclusions regarding O3-

32 related effects on forest tree growth, productivity and carbon storage. Information supporting 

33 these previous conclusions comes from a variety of different types of studies and which cover an 

34 array of different species, endpoints and exposure methods. One subset of studies focused on 
5-8 

1 underlying mechanisms as they relate to growth, productivity and carbon storage including: 

2 reduced carbon dioxide uptake due to stomatal closure (U.S. EPA 2013, section 9.3.2.1); the 

3 upregulation of genes associated with plant defense, signaling, hormone synthesis and secondary 

4 metabolism (U.S. EPA 2013, section 9.3.3.2); the down regulation of genes related to 

5 photosynthesis and general metabolism (U.S. EPA 2013, section 9.3.3.2); loss of carbon 

6 assimilation capacity due to declines in the quantity and activity of key proteins and enzymes 

7 (U.S. EPA, 2013, section 9.3.5.1); and negative impacts on the efficiency of the photosynthetic 

8 light reactions (U.S. EPA, 2013, section 9.3.5.1). These new studies “have increased knowledge 

9 of the molecular, biochemical and cellular mechanisms occurring in plants in response to O3”, 

10 adding “to the understanding of the basic biology of how plants are affected by oxidative 

11 stress...” (U.S. EPA, 2013, p. 9-11). 

12 The recent studies cover a variety of exposure methods, species, and settings. In 

13 particular, a recent meta-analysis indicates a relationship between O3 concentrations in the 

14 northern hemisphere and effects with the potential to affect growth (i.e., stomatal conductance 

15 and photosynthesis) (U.S. EPA, 2013, section 9.4.3.1; Wittig et al., 2007). A second meta16 analysis, which quantitatively compiled peer-reviewed studies from the past 40 years, found that 

17 ambient concentrations reported in those studies significantly decreased annual total biomass 

18 growth (7%) across the studies (U.S. EPA, 2013, section 9.4.3.1). The ISA states that these two 

19 meta-analyses demonstrate the coherence of O3 effects on plant photosynthesis and growth 

20 across numerous studies and species using a variety of experimental techniques”, and including a 

21 recent study, that “recent meta-analyses have generally indicated that O3 reduced carbon 

22 allocation to roots (Wittig et al., 2009; Grantz et al., 2006)” (U.S. EPA, 2013, pp. 9-45 to 9-46). 

23 Recent field-based studies also have added to the evidence base. For example, a study 

24 conducted in forest stands in the southern Appalachian Mountains found that “the cumulative 

25 effects of ambient levels of O3 decreased seasonal stem growth by 30-50% for most tree species 

26 in a high O3 year in comparison to a low O3 year (McLaughlin et al., 2007a). The authors also 

27 reported that high ambient O3 concentrations can increase whole-tree water use and in turn 

28 reduce late-season streamflow (McLaughlin et al., 2007b)” (U.S. EPA, 2013, p. 9-43). 

29 Additionally, a recent study has provided concentration-response information for tree seedlings 

30 of an additional species beyond the 11 previously studied (Gregg, et al., 2003). This study on 

31 cottonwood expands the dataset of studied species such that the 12 species for which C-R 

32 functions are available in this review include deciduous, coniferous, eastern, western, sensitive 

33 and tolerant species (U.S. EPA 2013, section 9.6.2; U.S. EPA, 2014, section 6.2, Figure 6-2, 

34 Table 6-1).

35 The “previous O3 AQCDs concluded that there is strong evidence that exposures to O3

36 decreases photosynthesis and growth in numerous plant species” and that “[s]tudies published 
5-9 

1 since the 2008 review support those conclusions” (U.S. EPA, 2013, p. 9-42). The previously 

2 available strong evidence included the development of robust concentration-response (C-R) 

3 functions for tree seedling biomass loss in 11 species under the National Health and 

4 Environmental Effects Research Laboratory-Western Ecology Division (NHEERL-WED) 

5 program. This series of experiments used open-top-chambers (OTC) to study seedling growth 

6 response under a variety of O3 exposures (ranging from near background to well above current 

7 ambient concentrations) and growing conditions (U.S. EPA 2013, section 9.6.2, Lee and Hogsett, 

1996). 8 

9 We additionally recognize that, because trees are long lived, in addition to the effects of 

10 O3 exposures over the annual growing season, trees and other perennials can also cumulate 

11 effects across multiple years. For example, growth affected by a reduction in carbohydrate 

12 storage in one year may result in the limitation of growth in the following year, so that effects 

13 “carry over” from one year to another (U.S. EPA, 2013, section 9.4.8; Andersen, et al., 1997). In 

14 past reviews, such carry-over effects have been documented in the growth of some tree seedlings 

15 and in roots. A number of recent studies based on the FACE exposure method in a planted forest 

16 at the Aspen FACE site in Wisconsin have augmented and supported the earlier information. 

17 These studies observed tree growth responses over seven years beyond the seedling growth stage 

18 and growing in field settings more similar to natural forest stands than OTC studies. In addition

19 to affecting tree heights, diameters, and main stem volumes in the aspen community, elevated O3

20 was reported to change intra- and inter-species competition (Kubiske et al., 2006; Kubiske et al., 

21 2007). For example, O3 treatments increased the rate of conversion from a mixed aspen-birch 

22 community to a birch dominated community, potentially changing intra- and inter-species 

23 competition. 

24 The EPA comparison of biomass results from the first seven years of the recent study by 

25 Kubiske et al (2007) to that predicted using the C-R function established from the earlier OTC 

26 experiments indicated close agreement (U.S. EPA 2013, Section 9.6.3). Accordingly, the ISA 

27 concludes that “[o]verall, the studies at the Aspen FACE experiment were consistent with many 

28 of the open-top chamber (OTC) studies that were the foundation of previous O3 NAAQS 

29 reviews” and that “[t]hese {recent} results strengthen the understanding of O3 effects on forests

30 and demonstrate the relevance of the knowledge gained from trees grown in OTC studies” (U.S. 

31 EPA 2013, p. 2-38). In addition to growth effects, these scientists also found that elevated O3

32 decreased birch seed weight, germination, and bud starch levels as well as aspen bud size. The 

33 effects on birch seeds could lead to a negative impact on species regeneration, while the bud 

34 effects may have been related to the observed delay in spring leaf development and have the 

35 potential to alter carbon metabolism of overwintering buds. These latter effects likely have 

36 implications for the subsequent growing season (i.e. carry-over effects) in the following year, 
5-10 

1 including effects on forest biomass, buds and seeds that carry into subsequent years (U.S. EPA, 

2 2013, section 9.4). 

3 The recent studies, in combination with the entire body of evidence, form the basis for 

4 the ISA determinations that: 1) there is a causal relationship between O3 exposures and reduced 

5 vegetative growth; 2) there is a causal relationship between O3 exposures and biomass 

6 accumulation, including altered carbon allocation to below ground tissues, rates of leaf and root 

7 production, and turnover and decomposition that can alter below-ground biogeochemical cycles; 

8 and 3) there is likely to be a causal relationship between O3 exposure and a reduction in carbon 

9 sequestration in terrestrial ecosystems (U.S. EPA, 2013, Table 2-2). Therefore, the current 

10 evidence base substantiates prior conclusions of O3-related effects on tree growth, productivity 

11 and carbon storage. The results from some of these studies are discussed more fully under 

12 different questions below. 

13 • To what extent have important uncertainties in the evidence identified in the last 

14 review been reduced and/or new uncertainties emerged?

15 As stated above, the ISA concludes that the new evidence confirms, strengthens or 

16 expands our understanding of O3 effects on plants. Much of this new evidence is focused on the 

17 molecular and genetic level, providing very important new mechanistic information that in some 

18 cases enhances our understanding of the complexity of the O3–plant response. This information 

19 has, in general, reduced overall uncertainties at the subcellular and cellular scales. However, 

20 because these studies were primarily conducted using artificial exposure conditions and model 

21 plants, uncertainties remain regarding the extent to which these plant responses reflect those in 

22 other plant species and exposure conditions (U.S. EPA, 2013, section 9.3.6). With regard to O3

23 impacts at the whole plant, species, and ecosystem scales, recent information has informed our 

24 understanding of associated uncertainties in a variety of ways. 

25 Importantly, one key uncertainty has been significantly reduced. This relates to the C-R 

26 functions we have used in previous reviews to estimate tree seedling biomass loss for different 

27 O3 exposure conditions (U.S. EPA, section 9.6). As these functions were derived from OTC 

28 experiments, an associated uncertainty has been with regard to the extent to which they reflected 

29 concentration-response relationships for field conditions. In the current review, EPA staff have 

30 conducted an analysis comparing OTC data with field-based data for one crop and one tree 

31 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)2 32 and field-based data (Soy FACE), 

 

2 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 
5-11 

1 as discussed in section 5.3 below. The second was done using aspen seedling OTC data from the 

2 NHEERL-WED and field-based data (Aspen FACE). The result of the aspen analysis, similar to 

3 that for soybean, showed very close agreement between the predictions based on NHEERL4 WED data and Aspen FACE observations, even when comparing the results of experiments that 

5 used different exposure methodologies, different genotypes, locations, and durations. Based on 

6 this analysis, the ISA additionally stated that “the [C-R] function based on one year of growth 

7 was applicable to subsequent years” (of the six-year dataset) (U.S. EPA, 2013, section 9.6.3.2). 

8 This information reduces the uncertainties associated with potential impacts of other 

9 experimental factors on the O3-plant response. Other studies, such as the meta-analyses 

10 discussed in the ISA and below, also demonstrate the coherence of O3 growth effects across 

11 numerous studies and species that used a variety of experimental techniques. While recognizing 

12 that uncertainties may remain for some individual species for which the database is relatively 

13 less robust (such as the more recent information on cottonwood), taken together, this information

14 substantially reduces uncertainties associated with use of the tree seedling OTC-derived C-R 

15 functions to predict the response of trees beyond the seedling stage in field settings. Thus, in the 

16 current review, the ISA and WREA have continued to use these functions to estimate tree growth 

17 response for different cumulative O3 exposures (U.S. EPA, 2013, section 9.6.2; U.S. EPA, 2014, 

18 section 6.2). 

19 Other uncertainties associated with studying or modeling O3 impacts on trees, including 

20 those identified in the last review, still remain, due in part to a lack of additional research but 

21 also due to the growth characteristics of trees which present a unique set of experimental 

22 challenges. For example, while trees are long-lived, with life spans which range from decades to 

23 centuries, most studies are designed to take place within an annual or 2-3 year timeframe, which 

24 represents only a small fraction of the lifetime of a tree. Further, trees grow very large, making it 

25 difficult to use controlled exposure environments beyond the seedling or sapling growth stage. 

26 Additionally, sensitivity to O3 varies over the life of the tree so that different growth stages of the 

27 tree may be more or less sensitive and this variation in growth-stage sensitivity is species28 specific. Thus, while some limited information exists regarding tree sensitivity beyond the 

29 seedling growth stage (e.g. aspen, cottonwood) and in some species for both seedling and mature 

30 trees within a species (e.g., red oak), it remains uncertain to what degree effects observed on 

31 trees during one growth stage (e.g., seedling) can be extrapolated to trees at other growth stages. 

32 An analysis in the WREA comparing seedling to adult tree biomass loss, discussed in 5.2.2 

 

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).
5-12 

1 below, informs our consideration of this remaining uncertainty (U.S. EPA, 2014, section 

2 6.2.1.1). 

3 Lastly, we recognize that exposures in one year have the potential to cause effects not 

4 observed until a subsequent year (termed “carry over” effects). While recent studies provide 

5 evidence of some carry-over effects and the potentially serious implications they could have for 

6 associated ecosystems and services, the true extent of this effect is unclear because most studies 

7 on the effects of O3 on growth do not measure or take into account the possibility of carry-over 

8 effects in subsequent years. Additionally, uncertainties remain regarding the extent of 

9 compounding of growth effects across multiple years for different air quality conditions. These 

10 uncertainties affect our characterization of such impacts, particularly the quantitative aspects, for 

11 differing exposure scenarios. Therefore our characterization of subsequent growing season 

12 effects is uncertain, affecting our ability to fully describe impacts of observed annual biomass 

13 losses, particularly in quantitative terms. In section 5.2.2 below, the potential variation of 

14 growth effect compounding across multiple years is discussed further drawing on the WREA 

15 evaluation of this (U.S. EPA, 2014, section 6.2.1.4). Further discussion of these tree-related 

16 uncertainties that are relevant in the context of informing our understanding of the quantitative 

17 and qualitative exposure and risk results are discussed in the appropriate sections below. 

18 • What are the ecosystem services potentially affected by O3 effects on tree growth, 

19 productivity and carbon storage and to what extent are they important from a 

20 public welfare perspective?

21 A variety of ecosystem services are potentially affected by O3 impacts on tree growth, 

22 productivity and carbon storage. The ISA identifies as causal the relationship of O3 and reduced 

23 productivity in terrestrial ecosystems and alteration of below ground biogeochemical cycles. It 

24 further identifies as likely to be a causal relationship O3impacts on reduced carbon sequestration 

25 in terrestrial ecosystems; alteration of terrestrial ecosystem water cycling; alteration of terrestrial 

26 community composition (U.S. EPA, 2013, Table 9-19). These effects are important to the 

27 public welfare, and in particular include those associated with national parks, national refuges 

28 and other protected areas ranging to the harvesting of timber for commercial uses. The 

29 ecosystem services most directly affected by biomass loss include: (1) habitat provision for 

30 wildlife, (2) carbon storage, (3) provision of food and fiber, and (4) pollution removal (see also 

U.S. EPA, 2014, Figure 6-1).3 31 Less direct impacts can occur on process-related effects such as 

32 nutrient and hydrologic cycles. 

 

3 The impacts of O3 in ambient air on some of these services are characterized in the WREA, as described 

in section 5.2.2 below.
5-13 

1 With regard to the extent of O3 impact on carbon transport or allocation to roots, 

2 information from a few recent individual studies have shown mixed results, including negative, 

3 non-significant, and positive effects on root biomass and root:shoot ratio. However, as assessed 

4 in the ISA, “[g]enerally, there is clear evidence that O3 reduces carbon transport or allocation to 

5 roots”, although (U.S. EPA, 2013, p. 9-44). 

6 • To what extent does the available evidence indicate the occurrence of O3-related 

7 effects on forest growth, productivity and carbon storage attributable to 

8 cumulative exposures at lower ambient O3 concentrations than previously 

9 established or to exposures that might be expected to occur under the current 

10 standard? 

11 The evidence available in this review indicates that O3-induced effects on tree growth, 

12 productivity and carbon storage occur as a result of cumulative exposure concentrations similar 

13 to those identified in the previous review and that these effects can occur at exposures associated 

14 with air quality conditions that might be expected to meet the current standard. We first consider 

15 the evidence on O3 effects on growth, particularly for tree seedlings, and that supports the C-R 

16 functions describing the relationship between cumulative O3 exposure and reduction in growth 

17 (relative biomass loss). As described above, this evidence base currently includes functions for 

12 species.

4 18 Figure 5-1, below (based on species-specific composite functions described in the 

19 ISA and WREA), illustrates the appreciable variation in sensitivity across individual species 

20 (U.S. EPA 2013, section 9.6.2; U.S. EPA, 2014, section 6.2, Table 6-1 and Figure 6-2). For 

21 example, in seven or the 12 species, the W126 index value for which 2% seedling biomass loss is 

22 estimated is below 8 ppm-hours and in the other five species, the W126 value for which 2% 

23 biomass loss is estimated is above 18 ppm-hours. Within the group of seven more sensitive 

24 species, the most sensitive are cottonwood and black cherry (U.S. EPA 2014, section 6.2). 

 

4 Among the 12 species, in addition to the functions for 11 species studied in the FACE research, is the 

field study by Gregg et al (2003) for cottonwood. As noted in discussion of uncertainties above, there is only the 

single study available for this species which focused on a gradient of O3 exposure extending from the New York 

metropolitan area which may contribute uncertainty to the quantitative characterization of C-R.
5-14 

1 

2 Figure 5-1. Relative biomass loss in seedlings for 12 studied species in response to seasonal 

3 ozone concentrations in terms of seasonal W126 index values (U.S. EPA 2014, 

4 Figure 6-2).

5 In considering the evidence for the two most sensitive species, we recognize the more 

6 limited quantitative information for cottonwood. The C-R function for this species is based on 

7 the study by Gregg et al (2003) which described O3 effects on growth on eastern cottonwood 

8 (Populus deltoides) saplings along an O3 gradient between urban New York City and downwind 

9 rural areas for the years 1992, 1993 and 1994. While the authors additionally studied 

10 cottonwood response in a companion controlled OTC experiment, the C-R function described 

11 here is based on the field data and reflects the single study (U.S. EPA, 2013, sections 9.2.5 and 

12 9.6.3). Given these factors, as noted in consideration of uncertainties above, we recognize there 

13 may be additional uncertainty in this function. 

14 Using the 11 concentration-response (C-R) functions for tree seedlings from the FACE 

15 research, and with the addition of the cottonwood C-R function, Figure 5-2 below presents three 

16 approaches for characterizing the median response function (U.S. EPA, 2014, section 6.2.1.2 and 

17 Figure 6-4). The first approach used a composite function developed from the C-R functions for 

18 all 52 tree seedling studies (across the 12 species) combined (red line). The second approach 

19 used only the median C-R function, when available, for each of the 12 tree species included

(green line)5 20 . The third approach used a stochastic sampling method to randomly select a C-R 

 

5 For some species, only one study was conducted so that C-R function was used.

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
5-15 

1 function for each species from the studies conducted for that species. For some species only one 

2 study was available (e.g., Red Maple), and for other species there were as many as 11 studies 

3 available (Ponderosa Pine). The process was repeated 1,000 times (grey lines), and the median 

4 value was plotted for biomass loss values of 1% to 7%, and 10% (the bar associated with each

median point denotes the 25th and 75th 5 percentile values. This third approach provides an 

6 illustration of the effect of within-species variability on estimates of the median species. The 

7 median W126 values are similar, when using all of the studies or just the composite C-R function 

8 for each species; however, the median value is higher when within-species variability is included

9 (U.S. EPA, 2014, section 6.2.1.2). Among these three approaches, the median seasonal W126 

10 index value for which a two percent biomass loss is estimated in seedlings for the studied species 

11 ranges between approximately 7 and 14 ppm-hrs. 

12 

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

14 (From U.S. EPA 2014, Figure 6-4) 

15 We further consider several recent studies that have provided important evidence of 

16 growth effects occurring in the field at ambient exposure concentrations, including some 

17 exposure concentrations which are at or below the level of the current standard. For example, 

18 the study by McLaughlin et al., (2007a, b) investigated the effects of ambient O3 on tree growth 

19 (measured by change in stem circumference) in the field at 3 forest sites (Cade’s Cove, Look 
5-16 

1 Rock and Oak Ridge) in the southern Appalachian Mountains for three years (2001, 2002, 2003) 

2 (U.S. EPA, 2013, section 9.4.5). Compared to the growth observed in 2001 (base year), the 

3 authors reported that ambient O3 exposures, decreased seasonal stem growth by 30-50% for most 

4 tree species in a high O3 year (2002), relative to the 2001 baseline. As shown in Table 5-6 below 

5 (modified from Table 6-4 in U.S. EPA, 2014), the seasonal W126 index values ranged from 

6 approximately 9 to 20 ppm-hours in the low O3 year (2003) and from approximately 18 to 40 

7 ppm-hours in the high O3 year (2002), across the three sites. When these exposure 

concentrations for the three years are expressed in terms of the 4th 8 highest daily maximum 8-hour 

9 average concentration, the values range from 71 to 90 ppb during the low year (2003) to 82 – 

102 ppb during the high year (2002). The 3-year average of these 4th highest daily maximums6 10 

11 for the period 2001-2003 are 76 ppb, 93 ppb, and 87 ppb for Cade’s Cove, Look Rock, and Oak 

12 Ridge, respectively. Thus, while two of the years at Cade’s Cove had annual values at or below 

13 75 ppb, the concentrations at all sites for the 2001- 2003 period appear to exceed the current 

14 standard. Only one species, tulip poplar, has data for all three years and was present at all three 

15 sites (U.S. EPA, 2014, Table 6-4). This subset of data is what is shown in table 5-1 below. 

16 Table 5-1. Ozone concentrations associated with effects on tulip poplar in southern 

17 Appalachian Mountains (2001-2003). 

Ambient O3 concentrations for McLaughlin et al., 2007A

Cade’s Cove Look Rock Oak Ridge

Max4 

high

(ppm)

W126

(ppmhrs)

Percent change 

in 

circumferenceB

Max4 

high

(ppm)

W126

(ppmhrs)

Percent change 

in 

circumferenceB

Max4 

high

(ppm)

W126

(ppmhrs)

Percent change 

in 

circumferenceB

2001 75 15 baseline 86 23 baseline 85 20 baseline

2002 82 18 -62% 102 40 -26% 99 32 -49.6%

2003 71 9 N/A 90 20 -38% 78 11 7.5%

2001-

2003 76 N/A 93 28 87 21

AAir quality values obtained from http://www.epa.gov/ttn/airs/airsaqs/detaildata/downloadaqsdata.htm. W126 

rounded to nearest whole ppm-hr. 2002 and 2003 study results measured as % change in tulip poplar stem 

circumference compared to 2001 baseline.

B Study results for percent change in circumference of Tulip Poplar from WREA Table 6-4.

18 

19 Based on the table above, appreciable growth reductions occurred in tulip poplar at all 

20 three sites during a high O3 year (2002), though it spanned a rather large range of -26% to -62%. 

 

6 This is the form of the current standard. Thus, the current standard is met when the 3-year average of the 

4th highest daily maximum 8-hour concentrations is at or below 75 ppb.
5-17 

1 The results of this study are hard to interpret given the different exposure base lines, magnitude 

2 of changes in subsequent year exposures, and other undocumented site-specific factors. The 

3 study findings do show, however, that higher O3 concentrations (above the current standard) 

4 significantly reduced growth compared to that in a lower O3 year. 

5 With regard to evidence of effects associated with exposures that may be allowed by the 

6 current standard, we additionally consider the study by Gregg et al (2003) referenced above that 

7 documented O3-induced growth loss on eastern cottonwood (Populus deltoides) saplings along 

8 an O3 gradient between urban New York City and downwind rural areas for the years 1992, 1993 

9 and 1994 (U.S. EPA, 2013, section 9.2.5). Fourth highest daily max 8-hour average 

10 concentrations during the experimental period (July 7 – Sept 20) ranged from 0.056 ppm to 

11 0.098 ppm along the gradient (data in 2008 O3 docket) while W126 index values ranged from 1.9 

12 to 12 ppm-hrs. At some sites, the highest daily maximum 8-hour average ambient concentrations 

13 (0.056 ppm, 0.072 ppm, and 0.073 ppm) were below the level of the current standard, indicating 

14 that the standard would likely have been met. Figure 5-3 below shows the growth suppression 

15 response of the cottonwood saplings to the range of O3 exposures occurring along the gradient. 

16 

17 Figure 5-3. Association of O3 with cottonwood biomass in gradient downwind of New 

18 York City (Gregg et al., 2003). Symbol Key: Squares (1992); Circles (1993); 

19 Triangles (1994); Shaded (urban); open (rural). Figure modified from Gregg et al., 

20 2003 and U.S. EPA, 2007 Figure 7-17 to add 8-hour average concentrations (red 

21 font).
5-18 

1 We additionally recognize some limitations in the extent to which some studies that look 

2 at O3-induced effects on tree growth, productivity and carbon storage can inform consideration 

3 of effects that might occur under air quality conditions that meet the current standard. In some 

4 cases this is because studies were designed to look only at what is happening under ambient 

5 conditions and/or above ambient conditions and the ambient conditions often did not meet the 

6 standard. In addition, because tree growth is known to be affected by cumulative exposures, 

7 studies do not commonly provide exposure concentration data pertaining to daily 8-hour 

8 averages. Unless hourly O3 concentration data are provided, translation from one index to 

9 another is not usually feasible. Thus, to consider effects that might occur under conditions that 

10 meet the current standard we have traditionally relied on monitored hourly O3 air quality data to 

11 calculate both types of exposure metrics (i.e. cumulative and 8-hour average) used in 

12 combination with the robust tree seedling C-R functions developed in OTC studies (as described 

13 above) to predict the magnitude of the growth effects associated with exposure concentrations at 

14 or below that of the current standard. These tree seedling C-R functions are also used in the 

15 WREA estimates of effects on tree growth that might be expected to remain under air quality 

16 conditions that just meet the current secondary standard (discussed in section 5.2.2 below). 

17 • To what extent does currently available evidence suggest locations where the 

18 vulnerability of sensitive species, ecosystems and/or their associated services to 

19 O3-related effects on forest tree growth, productivity and carbon storage would 

20 have special significance to the public welfare?

21 Areas with special significance to the public welfare include federally designated Class I, 

22 non-Class I national parks, and other areas set aside to provide similar public welfare benefits. 

23 Therefore Table 5-2 provides some examples of Class I sites where the current secondary 

24 standard is met but W126 index values fall above 15 ppm-hours. At W126 index values above 

25 15 ppm-hours, relative biomass loss in seedlings might be expected to range from less than 1% 

26 to more than 30% across the 12 species for which C-R functions have been developed (as 

27 described above). Relative biomass loss above 5% might be expected in seedlings of six of the 

28 12 species and above 2% in seedlings (but less than 5%) in a seventh, indicating that if present in 

29 these specially protected areas, meeting the current standard may not prevent W126 index values 

30 that may cause these levels of growth effect. 

31 

32 
5-19 

1 Table 5-2. Examples of counties containing Class I areas where recent air quality might be 

2 expected to meet the current standard and where 3-yr W126 index values are 

3 above 15 ppm-hrs. 

Monitor ID Years 3-yr 8-

hr max

3-yr W126

(annual values)

State/Co. Class I Area

806710041 2008-2010 71 15

(16, 11, 18)

Colorado/ La Plata Weminuche

WAa

806710041 2009-2011 74 17

(11, 18, 21)

Colorado/ La Plata Weminuche WA

806710041 2010-2012 73 19

(18, 21, 18)

Colorado/ La Plata Weminuche WA

80830101 2006-2008 71 18

(24, 18, 14)

Colorado/

Montezuma

Mesa Verde NPb

80830101 2007-2009 69 16

(18, 14, 15)

Colorado/

Montezuma

Mesa Verde NP

4005800 2006-2008 70 19

(22,19, 22)

Arizona/ Coconino Grand Canyon 

NP

4005800 2007-2009 68 15

(19, 22, 11)

Arizona/ Coconino Grand Canyon 

NP

4005800 2010-2012 72 18

(15, 18, 20)

Arizona/ Coconino Grand Canyon 

NP

49053013 2006-2008 71 21

(24, 19, 20)

Utah/ Washington Zion NP

49053013 2007-2009 70 18

(19, 20, 15)

Utah/ Washington Zion NP

49053013 2008-2010 70 18

(20, 15, 21)

Utah/ Washington Zion NP

49053013 2009-2011 70 18

(15, 21,18)

Utah/ Washington Zion NP

49053013 2010-2012 73 20

(21, 18, 22)

Utah/ Washington Zion NP

49037010 2006-2008 71 18

(19, 18, 18)

Utah/ San Juan Canyonlands 

NP

49037010 2007-2009 70 16

(18, 18, 13)

Utah/ San Juan Canyonlands 

NP

49037010 2010-2012 69 15

(14, 14, 17)

Utah/ San Juan Canyonlands 

NP

Based on data from http://www.epa.gov/ttn/airs/airsaqs/detaildata/downloadaqsdata.htm

a = Wilderness Area; b= National Park

4 
5-20 

1 5.2.2 Exposure/Risk-based Considerations

2 The WREA presents a number of quantitative analyses of cumulative exposure and risk 

3 related to tree growth, productivity and carbon storage for air quality scenarios intended to 

4 inform our consideration of exposure and risk for considerations associated with the current 

5 standard (see Table 5-3 below, U.S. EPA, 2014, chapter 6). 

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

7 productivity and carbon storage. 

Species Level 

Effects

Ecosystem Level 

Effects Ecosystem Services

WREA 

estimatesA

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 6-11)

• Carbon storage, nationally (WREA, Table 6-18) 

• Carbon storage, in 5 urban areas (WREA, Table 

6-20) 

• Air pollutant removal in 5 urban areas (WREA, 

Table 6-21) 

A See WREA chapter 6. BThe total geographic area includes the contiguous U.S..

8 

9 The relevant quantitative exposure and risk analyses for tree biomass loss, productivity 

10 and carbon storage include: 

11 1) Species-specific and composite biomass loss estimates. 

12 2) National-scale assessments for: a) basal area weighted relative biomass loss for tree 

13 seedlings; b) timber production; c) carbon sequestration. 

14 3) Case study-scale assessments for: a) carbon sequestration; b) air pollution removal. 

15 • For what air quality scenarios were exposures and risks estimated? What

16 approaches were used to estimate W126 exposures for those conditions? What 

17 are associated limitations and uncertainties?

18 In the analyses related to O3 effects on tree growth, productivity and carbon storage, 

19 quantitative estimates were developed for five air quality scenarios by the methodology 

20 summarized in Table 5-4 below. In general, this methodology involved two steps. The first is 

21 derivation of the average W126 value (across the three years) at each monitor location. This 

22 value is based on unadjusted data for recent conditions and model-adjusted concentrations for the 
5-21 

1 four other scenarios. The development of model adjusted concentrations was done for each of 9 

2 regions independently (see U.S. EPA, 2014, section 4.3.4.1). In the second step, national-scale 

3 spatial surfaces (W126 values for each model grid cell) were created using the monitor-location 

4 values and the Voronoi Neighbor Averaging (VNA) spatial interpolation technique (details on 

5 the VNA technique are presented in U.S. EPA, 2014 Appendix 4A). 

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

7 values was derived for each air quality scenario.

Scenario

Development of W126 values for Each Air Quality Scenario

Monitor-location-specific calculations

and any model 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 values to create a 

national distribution of W126 

values within model grid-cells 

for each scenario.

Current Standard

2006-2008 hourly O3 concentrations at each monitor 

location are model-adjustedA to create a three year 

record of O3 concentrations that just meets the 

current standard (see WREA, section 4.3.4).

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

Hourly O3 concentrations at each monitor location, 

within each modeling region, are model-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 locationspecific 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 4 and 7, and Appendices 

4A and 7A). 
5-22 

1 

2 During the recent conditions period (2006 through 2008), the average W126 index values 

3 (across the three-year recent conditions period) at the monitor locations ranged from below 10 

4 ppm-hrs to 48.6 ppm-hrs (U.S. EPA 2014, Figure 4-4 and Table 4-1). Across the nine modeling 

5 regions, the maximum average W126 index values ranged from 48.6 ppm-hrs in the west region 

6 down to 6.6 ppm-hrs in the northwest region. After model-adjustment of the 2006-2008 data to 

7 just meet the current standard in each region, the region-specific maximum values range from 

8 18.9 ppm-hrs in the west region to 2.6 ppm-hrs in the northeast region (U.S. EPA, 2014, Table 4-

9 1). After application of the VNA technique to the current standard scenario monitor location 

10 values, the average W126 values was below 15 ppm-hrs across the national surface with the 

11 exception of a very small area of the southwest region (near Phoenix) where the average W126 

12 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 values at a 12x12 km2 13 grid 

14 resolution rather that at the exact location of a monitor. This indicates one uncertainty associated 

15 with this aspect of the approach to estimating W126 values for the model-adjusted air quality just 

16 meeting the current standard.

17 The WREA also recognizes other sources of uncertainty for the W126 estimates for each 

18 air quality scenario and qualitatively characterizes the magnitude of uncertainty and potential for 

19 directional bias. These sources of uncertainty are described in more detail in the WREA Chapter 

20 4 and summarized below. Because the W126 estimates generated in the air quality analyses are 

21 inputs to the vegetation risk analyses for biomass loss and foliar injury, any uncertainties in the 

22 air quality analyses are propagated into the those analyses (U.S. EPA 2014, section 8.5). 

23 An important large uncertainty in the analyses is the assumed response of the W126 

24 concentrations to emissions reductions needed to meet the existing standard (U.S. EPA, 2014, 

25 section 8.5.1). We note that any approach to characterizing O3 air quality over broad geographic 

26 areas based on concentrations at monitor locations will convey inherent uncertainty. The model27 based adjustments, based on U.S.-wide emissions reductions in oxides of nitrogen (NOx), do not 

28 represent air quality distributions from an optimized control scenario that just meets the current 

29 standard (or target W126 values for other scenarios), but characterize one potential distribution 

30 of air quality across a region when all monitor locations meet the standard (U.S. EPA 2014, 
5-23 

section 4.3.4.2).7 1 An additional uncertainty comes from the creation of a national W126 surface 

2 using the VNA technique to interpolate recent air quality measurements of O3. In general, spatial 

3 interpolation techniques perform better in areas where the O3 monitoring network is denser. 

4 Therefore, the W126 estimated in the rural areas in the West, Northwest, Southwest, and West 

5 North Central with few or no monitors (Figure 2-1) are more uncertain than those estimated for 

areas with more dense monitoring8 6 . Additionally, the surface is created from the three-year 

7 average at the monitor locations, rather than creating a surface for each year and then averaging 

8 across years at each grid cell; the potential impact of this on the resultant estimates is considered 

9 in the WREA (U.S. EPA, 2014, Appendix 4A). 

10 • What are the nature and magnitude of exposure- and risk-related estimates for 

11 tree growth, productivity, and carbon storage under recent conditions or 

12 conditions remaining upon meeting the current standard? To what extent are 

13 these exposures and risks important from a public welfare perspective?

14 The WREA used the C-R functions for 12 species described above with information on 

15 the distribution of those species across the U.S., and average W126 exposure estimates to 

16 estimate relative biomass loss for each of the studied species for each national air quality 

17 scenario (U.S. EPA, 2014, section 6.2.1.3 and Appendix 6A). For example, the estimates of 

18 relative biomass loss of Ponderosa Pine for air quality model-adjusted to just meet the current 

19 standard are illustrated in Figure 5-4 below. While relative biomass loss below 2% is estimated 

20 for most areas where this species is found, estimates in some areas of the southwest fall in a bin 

21 defined as 2.01-4% biomass loss (U.S. EPA 2014, Figure 6-7 and Appendix 6A). 

22 

23 

24 

25 

26 

27 

 

7 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-ozone urban monitors for reductions. However, the assumption of broad regional or 

national NOx reductions are not unreasonable given current EPA regulations such as the Clean Air Interstate Rule 

(CAIR), which requires NOx cuts across the Eastern U.S. to reduce regional ozone transport, 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.
5-24 

1 

2 Figure 5-4. Relative biomass loss of Ponderosa Pine for air quality model-adjusted to just 

3 meet the current standard (U.S. EPA 2014, Figure 6-7).

4 The WREA also developed national-scale estimates of O3 biomass ecosystem-level 

5 impacts considering the 12 studied species together (U.S. EPA 2014, section 6.8, Table 6-22). 

6 This was done using the species-specific biomass loss C-R functions, information on prevalence 

7 of the studied species across the U.S., and a weighting approach based on proportion of the basal 

8 area that each species contributes. The RBL values for multiple tree species were weighted by 

9 their basal area and combined into a weighted RBL value. The weighted relative biomass loss 

10 (wRBL) is intended to inform our understanding of the potential magnitude of the ecological 

11 effect that could occur in some ecosystems. Specifically, the more basal area that is affected in a 

12 given ecosystem, the larger the potential ecological effect. A wRBL value for each grid cell is 

13 generated by weighting the RBL value for each studied tree species found within that grid cell by 

14 the proportion of basal area it contributes to the total basal area of all tree species within the grid 

15 cell, and then summing those individual wRBLs. The percent of total basal area that exceeds a 

16 2% weighted relative biomass loss in the recent conditions scenario is 10.1% (U.S. EPA 2014, 

17 Table 6-24). Based on the average W126 values estimated for the air quality scenario just 
5-25 

1 meeting the current standard across the contiguous U.S., the WREA estimates 0.8 % of the total 

2 geographic area to have a wRBL above 2% (U.S. EPA 2014, Table 6-24). 

3 We also consider WREA estimates (quantitative and qualitative) of effects on several 

4 ecosystem services. First, impacts on growth related to O3 concentrations in federally-designated 

5 Class I areas were derived from an average weighted RBL value (discussed above) for 145 of the 

6 155 Class I areas (U.S. EPA 2014, section 6.8.1). Given established objectives for Class I areas 

7 (e.g., to maintain in perpetuity), effects in Class I areas may be considered to have the potential 

8 to adversely affect the intended use of the ecosystem, i.e., to maintain in pristine natural 

9 conditions for future generations. For the recent conditions scenario, this analysis estimates 

10 wRBL values above 2% in 13 of the 145 assessed Class I areas. In comparison, the analysis 

11 estimates wRBL values above 2% in only 2 areas based on average W126 values estimated for 

12 the current standard scenario (U.S. EPA 2013, Table 6-26). 

13 The WREA also presents national-scale estimates of the effects of biomass loss on timber 

14 production and agricultural harvesting, as well as on carbon sequestration. The WREA used the 

15 O3 C-R functions for tree seedlings to calculate relative biomass loss. Because the forestry and 

16 agriculture sectors are related, and trade-offs occur between the sectors, the WREA also 

17 calculated the resulting market-based welfare effects of O3 exposure in the forestry and 

18 agriculture sectors. In the analyses for commercial timber production, based on the average 

19 W126 values estimated for the air quality scenario just meeting the existing standard, relative 

20 biomass losses (RBL) estimates were below one percent in all regions except the Southwest, 

21 Southeast, Central, and South regions (U.S. EPA, 2014, section 6.3, Table 6-8) (see U.S. EPA, 

22 2014, Table 6-14 for clarification on region names) ). Relative biomass losses remain above one 

23 percent for the average W126 scenarios for 15 and 11 ppm-hrs in parts of the Southeast, Central, 

24 and South regions, and for the 7 ppm-hr scenario in the Southeast and South regions (U.S. EPA, 

25 2014, section 6.3, Table 6-8).

26 In addition to estimating changes in forestry and agricultural yields, the WREA presents 

27 estimated changes in consumer and producer/farmer surplus associated with the change in yields. 

28 Changes in biomass affect individual tree species, but the overall effect on forest ecosystem 

29 productivity depends on the composition of forest stands and the relative sensitivity of trees 

30 within those stands. Economic welfare impacts resulting from just meeting the existing and 

31 alternative standards were largely similar between the forestry and agricultural sectors --

32 consumer surplus, or consumer gains, generally increased in both sectors because higher 

33 productivity under lower O3 concentrations increased total yields and reduced market prices. 

34 Comparisons are not straightforward to interpret due to market dynamics. For example, because 

35 demand for most forestry and agricultural commodities is not highly responsive to changes in 
5-26 

1 price, there were more examples for which producer surplus, or producer gains declines. In 

some cases, lower prices reduce producer gains more than can be offset by higher yields.9 2 In 

3 general, the WREA estimates benefits for agricultural yield of approximately 1 billion dollars 

4 from differences in average W126 index values representing recent conditions from those 

5 representing model-adjusted air quality for the current standard (U.S. EPA 2014, section 6.3). 

6 The national-scale analysis of carbon dioxide (CO2) sequestration estimates substantially more 

7 storage under the current standard compared to recent conditions (U.S. EPA 2014, section 6.6.1, 

8 Table 6-18, Appendix 6B). In considering the significance of the potential climate and 

9 ecosystem service impact, we also take note of the large uncertainties associated with this 

10 analysis (see U.S. EPA 2014, Table 6-27). 

11 We additionally consider the WREA estimates of tree growth and ecosystem services 

12 provided by urban trees over a 25-year period, for five urban areas based on case-study scale 

13 analyses that quantified the effects of biomass loss on carbon sequestration and pollution 

removal (U.S. EPA 2014, sections 6.6.2 and 6.7).10 14 The urban areas included in this analysis 

15 represent diverse geography in the Northeast, Southeast, and Central regions, although do not 

16 include an urban area in the western part of the U.S. Estimates of the effects of O3-related 

17 biomass loss on carbon sequestration indicate the potential for an increase of somewhat more 

18 than a million metric tons of CO2 equivalents for average W126 values associated with meeting 

19 for the current standard scenario as compared to recent conditions. Somewhat smaller increases 

20 are estimated for the three W126 scenarios in comparison to the current standard scenario (U.S. 

21 EPA 2014, section 6.6.2 and Appendix 6D). 

22 In addition to the quantitative assessments discussed above, qualitative assessments for 

23 some ecosystem services, such as commercial non-timber forest products and recreation (U.S. 

24 EPA, 2014, section 6.4), aesthetic and non-use values (U.S. EPA, 2014, section 6.4), increased 

25 susceptibility to insect attack and fire damage (U.S. EPA, 2014, sections 5.3 and 5.4, 

26 respectively), were also conducted. Other ecological effects that are causally or likely causally 

27 associated with O3 exposure such as terrestrial productivity, water cycle, biogeochemical cycle, 

28 and community composition (U.S. EPA 2013, Table 9-19) were not directly addressed in the 

29 WREA due to a lack of sufficient quantitative information. 

 

9 The relative biomass loss estimates were also used with 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 multi-period, multi-commodity results over 60 to 100 years in 5-year time intervals 

when running the combined forest-agriculture version of the model. See Chapter 6, Section 6.3 of the WREA for a 

discussion of economic welfare and consumer and producer surplus.

10 The model, iTree, a peer-reviewed suite of software tools provided by USFS was used in this analysis.
5-27 

1 There is substantial heterogeneity in plant responses to O3, both within species, between 

2 species, and across regions of the U.S. The O3-sensitive tree species are different in the eastern 

3 and western U.S. -- the eastern U.S. has far more species. Ozone exposure and risk is somewhat 

4 easier to assess in the eastern U.S. because of the availability of more data and the greater 

5 number of species to analyze. In addition, there are more O3 monitors in the eastern U.S. but 

6 fewer national parks (U.S. EPA, 2014, chapter 8). 

7 • What are the uncertainties associated with both quantitative and qualitative

8 information?

9 Several key limitations and uncertainties, which may have a large impact on both overall 

10 confidence and confidence in individual analyses, are discussed here. Key uncertainties 

11 associated with the assessment of impacts on ecosystem services at the national and case-study 

12 scales, as well as across species, U.S. geographic regions and future years include those 

13 associated with the following seedling C-R functions, as well as interpolated and model-adjusted 

14 O3 concentrations used to estimate W126 exposures in the REA air quality scenarios. The 

15 uncertainties in the W126 exposure estimates are discussed above at the beginning of section 

16 5.2.2. 

17 With regard to the seedling C-R functions, the description of Figure 5-2 above provides 

18 some characterization of the variability of individual study results and the impact of that on 

19 estimates of W126 index values that might elicit different percentages of biomass loss in tree 

20 seedlings (U.S. EPA, 2014, section 6.2.1.2). Even though the evidence shows that there are 

21 additional species adversely affected by O3-related biomass loss, the WREA only has C-R 

22 functions available to quantify this loss for 12 tree species and 10 crop species. This absence of 

23 information only allows a partial characterization of the O3-related biomass loss impacts in trees 

24 and crops associated with recent O3 index values and with just meeting the existing and 

25 potential alternative secondary standards. In addition, there are uncertainties inherent in these C26 R functions, including the extrapolation of relative biomass loss rates from tree seedlings to 

27 adult trees and information regarding within-species variability. The overall confidence in the 

28 C-R function varies by species based on the number of studies available for that species. Some 

29 species have low within-species variability (e.g., many agricultural crops) and high 

30 seedling/adult comparability (e.g., Aspen), while other species do not (e.g., Black Cherry). The 

31 uncertainties in the C-R functions for biomass loss and in the air quality analyses are propagated 

32 into the analysis of the impact of biomass loss on ecosystem services, including provisioning 

33 and regulating services (U.S. EPA, 2014, Table 6-27). The WREA characterizes the direction 

34 of potential influence of C-R function uncertainty as unknown, yet its magnitude as high, 
5-28 

1 concluding that further studies are needed to determine how accurately the assessed species 

2 reflect the larger suite of O3-sensitive tree species in the U.S. (U.S. EPA, 2014, Table 6-27).

3 Another uncertainty associated with interpretation of the WREA biomass loss-related 

4 estimates concerns the potential for underestimation of compounding of growth effects across 

5 multiple years of varying concentrations. Though tree biomass loss impacts were estimated 

6 using air quality scenarios of 3-year average W126 values, the WREA also conducted an 

7 analysis to compare the impact of using a variable compounding rate based on yearly variations 

8 in W126 exposures to that of using a W126 index value averaged across three years. The WREA

9 compared the compounded values for two examples. In both examples, one species (Tulip Polar 

10 and Ponderosa Pine) and one climate region where that species occurred (Southeast and 

11 Southwest regions) were chosen and air quality values associated with just meeting the existing 

12 standard of 75 ppb were used. Within each region the WREA analysis used both the W126 value 

13 at each monitor in the region for each year and the three-year average W126 value using the 

14 method described in Chapter 4. The results show that the use of the three-year average W126 

15 index value may underestimate RBL values slightly, but the approach does not account for 

16 moisture levels or other environmental factors that could affect biomass loss (U.S. EPA, 2014, 

17 section 6.2.1.4 and Figure 6-13). In considering these results, we note that in both regions and in 

18 all three years, the three-year average W126 value is sometimes above and sometimes below the 

19 individual year W126 index value. 

20 In the national-scale analyses of timber production, agricultural harvesting, and carbon 

21 sequestration, the WREA used the FASOMGHG model, which includes functions for carbon 

22 sequestration, assumptions regarding proxy species, and non-W126 C-R functions for three

23 crops. However, FASOMGHG does not include agriculture and forestry on public lands, 

24 changes in exports due to O3 into international trade projections, or forest adaptation. Despite 

25 the inherent limitations and uncertainties, the WREA concludes that the FASOMGHG model 

26 reflects reasonable and appropriate assumptions for a national-scale assessment of changes in 

27 the agricultural and forestry sectors due to changes in vegetation biomass associated with O3

28 exposure (U.S. EPA, 2014, sections 6.3, 6.5, 6.6, and 8.5.2). 

29 In the case study analyses of five urban areas, the WREA used the iTree model, which 

30 includes an urban tree inventory for each area and species-specific pollution removal and carbon 

31 sequestration functions. However, iTree does not account for the potential additional VOC 

32 emissions from tree growth, which could contribute to O3 formation. Despite the inherent 

33 limitations and uncertainties, the WREA concludes that the iTree model reflects reasonable and 
5-29 

1 appropriate assumptions for a case study assessment of pollution removal and carbon 

2 sequestration for changes in biomass associated with O3 exposure (U.S. EPA, 2014, sections 

3 6.6.2, 6.7, and 8.5.2). 

4 The overall effect of the combined set of uncertainties on confidence in the interpretation 

5 of the results of the analyses is difficult to quantify. Due to differences in available information, 

6 the degree to which each analysis was able to incorporate quantitative assessments of uncertainty 

7 differed. 

8 5.3 CROP YIELD LOSS

9 This section considers the current evidence and exposure/risk information to inform 

10 consideration of the adequacy of the protection provided by the current standard from known and 

11 anticipated adverse welfare effects of O3 related to crop yield and other associated effects. Crops 

12 are important from a public welfare perspective because they provide food and fiber services to 

13 humans. This section includes a discussion of the policy-relevant science and weight-of14 evidence conclusions discussed in the ISA (section 5.3.1) and the exposure/risk results (section 

15 5.3.2) described in the second draft WREA. Important uncertainties and limitations in the 

16 available information are discussed throughout the sections. These discussions highlight the 

17 information we consider relevant to answering the overarching question and associated policy18 relevant questions included in this section. 

19 5.3.1 Evidence-based Considerations 

20 Ozone can interfere with carbon gain (photosynthesis) and allocation of carbon. As a 

21 result of decreased carbohydrate availability, fewer carbohydrates are available for plant growth, 

22 reproduction, and/or yield. For seed-bearing plants, these reproductive effects will culminate in 

23 reduced seed production or yield. The detrimental effect of O3 on crop production has been 

24 recognized since the 1960s, and current O3 concentrations in many areas across the U.S. are high 

25 enough to cause yield loss in a variety of agricultural crops including, but not limited to, 

26 soybeans, wheat, potatoes, watermelons, beans, turnips, onions, lettuces, and tomatoes. 

27 Increases in O3 concentration may further decrease yield in these sensitive crops while also 

28 causing yield losses in less sensitive crops (U.S. EPA 2013, section 9.4.4). The ISA concluded 

29 that the evidence is sufficient to determine that there is a causal relationship between O3

30 exposure and reduced yield and quality of agricultural crops (U.S. EPA 2013, Table 2-2).
5-30 

1 • To what extent has scientific information become available that alters or 

2 substantiates our prior conclusions regarding O3-related crop yield loss and of 

3 factors that influence associations between O3 levels and crop yield loss? 

4 In general, the vast majority of the new scientific information has substantiated our prior 

5 conclusions regarding O3 crop yield loss. On the whole, this evidence supports previous 

6 conclusions that exposure to O3 decreases growth and yield of crops. The ISA describes average 

7 yield loss reported across a number of meta-analytic studies have been published recently for 

8 soybean wheat, rice, semi-natural vegetation, potato, bean and barley (U.S. EPA 2013, section 

9 9.4.4.1). Meta-analysis allows for the objective development of a quantitative consensus of the 

10 effects of a treatment across a wide body of literature. Further, several new exposure studies 

11 continue to show decreasing yield and biomass in a variety of crops with increased O3 exposure 

12 (U.S. EPA 2013, section 9.4.4.1, Table 9-17). Research has linked increasing O3 concentration 

13 to decreased photosynthetic rates and accelerated aging (U.S. EPA 2013, section 9.4.4) in leaves, 

14 which are related to yield. Recent research has highlighted the effects of O3 on crop quality. 

15 Increasing O3 concentration decreases nutritive quality of grasses, decreases macro- and micro16 nutrient concentrations in fruits and vegetable crops (U.S. EPA 2013, section 9.4.4). The 

17 findings of these studies did not change our understanding of O3-related crop loss and little 

18 information has emerged on factors that influence associations between O3 levels and crop yield 

loss. 19 

20 • To what extent have important uncertainties identified in the last review been 

21 reduced and/or new uncertainties emerged? 

22 Important uncertainties have been reduced regarding the confidence placed in using crop 

23 exposure-response functions, especially for soybean. In general, the ISA reports consistent 

24 results across exposure techniques and across crop varieties. 

25 Two important uncertainties have been reduced regarding the C-R functions for yield 

26 effects of O3 in crop species, especially for soybean. First, in the last several reviews, the extent 

27 to which C-R functions developed in OTC predicted plant responses in the field and under 

28 different exposure conditions was not clear. In this review, staff from the EPA’s ORD/NCEA 

29 performed an analysis comparing OTC data with field-based data for one crop and one tree 

30 species (U.S. EPA, 2013, section 9.6.3.2). The crop comparison was done using soybean OTC 

31 data from NCLAN and field-based data from Soy FACE. The NCLAN studies were undertaken 

32 in the early to mid 1980's and provide the largest, most uniform database on the effects of O3 on 

33 agricultural crop yields (U.S. EPA 1996; U.S. EPA 2006; U.S. EPA 2013, sections 9.2, 9.4, and 
5-31 

9.6). 11 1 The Soy FACE experiment was a chamberless field-based exposure study in Illinois that 

2 was conducted from 2001 – 2009 (U.S. EPA 2013, section 9.2.4). Yield loss in soybean from O3

3 exposure at the Soy FACE field experiment was reliably predicted by soybean C-R functions 

4 developed in NCLAN (U.S. EPA, 2013, Section 9.6). This analysis supports the robustness and 

5 use of the C-R functions developed in NCLAN to predict relative yield loss to O3 exposure. 

6 A second area of uncertainty that was reduced is that regarding the application of the 

7 NCLAN C-R functions, developed in the 1980s, to more recent cultivars currently growing in the 

8 field. However, recent studies continue to find yield loss levels in crop species studied 

9 previously under NCLAN that reflect the earlier findings. There has been little new evidence 

10 that crops are becoming more tolerant of O3 (U.S. EPA, 2006a; U.S. EPA 2013). This is 

11 especially evident in the research on soybean. In a meta-analysis of 53 studies, Morgan et al. 

12 (2003) found consistent deleterious effects of O3 exposures on soybean from studies published 

13 between 1973 and 2001. Further, Betzelberger et al. (2010) has recently utilized the SoyFACE 

14 facility to compare the impact of elevated O3 concentrations across 10 soybean cultivars to 

15 investigate intraspecific variability of the O3 response. The C-R functions derived for these 10 

16 current cultivars were similar to the response functions derived from the NCLAN studies 

17 conducted in the 1980s (Heagle, 1989), suggesting there has not been any selection for increased 

18 tolerance to O3 in more recent cultivars. The 2013 ISA reported comparisons between yield 

19 predictions based on data from cultivars used in NCLAN studies, and yield data for modern 

20 cultivars from SoyFACE (U.S. EPA, 2013, section 9.6.3). They confirm that the average 

21 response of soybean yield to O3 exposure has not changed in current cultivars. Thus, staff 

22 concludes that at least for soybean, uncertainties associated with use of the NCLAN generated C23 R functions to estimate biomass loss in recent cultivars has been reduced.

24 • To what extent does the available evidence indicate the occurrence of O3-related 

25 effects on crop yield loss attributable to cumulative exposures at lower ambient O3

26 concentrations than previously established or to exposures at or below the level of 

27 the current standard? 

28 Very little evidence has emerged to indicate a lower level cumulative exposures that can 

29 affect crop yield (levels of concern), that have been based on C-R functions from OTC 

 

11 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. 
5-32 

1 experiments. As described above, the new evidence has strengthened the basis for using the 

2 information from the C-R functions. 

3 Where the current evidence on crop yield loss is not in terms of parts per billion 

4 concentrations over a specific exposure period such as eight hours, assessing whether O3

5 concentrations associated with meeting the current standard would allow crop yield effects is 

6 more complex. Where feasible, we have attempted to characterize the O3 exposures associated 

7 with crop yield loss in terms of seasonal W126 index and we have separately considered the 

8 extent to which such index values might be expected to occur in agricultural locations that meet 

9 the current standard. For example, Sedgwick and Sumner counties in Kansas met the level of the 

10 3-year 8-hr standard of 75 ppb in 2011. However, the W126 for 2011 in those counties was 19 

11 ppm-hours and would be predicted to result in a 9% yield loss for soybean grown in those 

12 counties. 

13 • To what extent does currently available evidence suggest locations where the 

14 vulnerability of sensitive species, ecosystems and/or their associated services to 

15 O3-related crop yield loss would have special significance to the public welfare?

16 During the previous NAAQS reviews, there were very few studies that estimated O3

17 impacts on crop yields at large geographical scales (i.e., regional, national or global). Recent 

18 modeling studies of the historical iimpact of O3 concentrations found that increased O3 generally 

19 reduced crop yield, but the impacts varied across regions and crop species (U.S. EPA, 2013, 

20 Section 9.4.4.1). The largest O3-induced crop yield losses were estimated to occur in high21 production areas exposed to elevated O3 concentrations, such as the Midwest and the Mississippi 

22 Valley regions of the United States. Among crop species, the estimated yield loss for wheat and 

23 soybean were higher than rice and maize. Additionally, satellite and ground-based O3

24 measurements have been used to assess yield loss caused by O3 over the continuous tri-state area 

25 of Illinois, Iowa, and Wisconsin. The results indicate that O3 concentrations during the assessed 

26 period reduced soybean yield, which correlates well with the previous results from FACE- and 

27 OTC-type experiments (U.S. EPA 2013, section 9.4.4.1). 

28 Thus, the recent scientific literature continues to support the conclusions of the 1996 and 

29 2006 Criteria Documents and 2013 ISA that ambient O3 concentrations can reduce the yield of 

30 major commodity crops in the U.S. and support the use of crop C-R functions based on OTC 

31 experiments. Agricultural areas that would be likely to have the most significance to the public 

32 welfare would be those high production areas for sensitive crops that also are exposed to high O3

33 concentrations, such as areas in the Midwest and Mississippi Valley regions. 
5-33 

1 5.3.2 Exposure/Risk-based Considerations

2 Two main analyses are conducted in the second draft WREA to estimate O3 impacts 

3 related to crop yield. Annual yield losses are estimated for 10 commodity crops and these 

4 estimates are then additionally used to estimate O3 impacts on producer and consumer economic 

5 surpluses (Table 5-5 below; U.S. EPA, 2014, sections 6.2, 6.5). 

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

Crop-level impactA Agri-Ecosystem ServicesB

REA estimates Annual Relative Yield Loss

Corn, Cotton, Potato, Sorghum, 

Soybean, Winter Wheat

Economic surplus to crop producers 

and consumers

ASee section 6.2 WREA. 3See section 6.5 WREA.

7 

8 • For what air quality scenarios were exposures and risks estimated? What

9 approaches were used to estimate W126 exposures for those conditions? What 

10 are associated limitations and uncertainties?

11 The WREA crop analyses described here were performed for five air quality scenarios by 

12 the methodology summarized in Table 5-4 above. In general, this methodology involved two 

13 steps. The first is derivation of the average W126 value (across the three years) at each monitor 

14 location. This value is based on unadjusted data for recent conditions and model-adjusted 

15 concentrations for the 4 other scenarios. The development of model adjusted concentrations was 

16 done for each of 9 regions independently (see U.S. EPA, 2014, section 4.3.4.1). In the second 

17 step, national-scale spatial surfaces (W126 values for each model grid cell) were created using

18 the monitor-location values and the Voronoi Neighbor Averaging (VNA) spatial interpolation 

19 technique (details on the VNA technique are presented in U.S. EPA, 2014 Appendix 4A). The 

20 results of model adjustments on estimated average W126 values and grid cell estimates produced 

21 from the VNA interpolation approach is summarized in section 5.2.2 above. A lowering of the 

22 highest values occurred with application of the interpolation method as a result of estimating 

W126 values at a 12x12 km2 23 grid resolution rather that at the exact location of a monitor. This 

24 indicates one uncertainty associated with this aspect of the approach to estimating W126 values 

25 for the model-adjusted air quality just meeting the current standard. Other areas of uncertainty

26 associated with the model adjustment and VNA interpolation approach are briefly summarized in 

27 section 5.2.2 above and described in more detail in the WREA (U.S. EPA, 2014, chapter 4 and 

28 Appendix 4A). 
5-34 

1 • What is the nature and magnitude of the cumulative exposure- and risk-related 

2 estimates for crop yield loss associated with remaining upon simulating just 

3 meeting the current O3 standard? What are the uncertainties associated this 

4 information? 

5 The WREA presents estimates of crop yield loss for the five air quality scenarios 

6 described above using C-R functions for commodity crops across the country (U.S. EPA, 2014, 

7 section 6.5). The largest reduction in O3 exposure-related crop yield loss occurs when moving 

8 from the recent conditions scenario to that for just meeting the current standard (U.S. EPA, 2014, 

9 section 6.5). In the analyses for agricultural harvest, the largest estimates of yield changes also 

10 occur when comparing the recent conditions scenario to that for the current standard. Under 

11 recent conditions, the West, Southwest, and Northeast regions generally have the highest yield 

losses. For the three average W126 scenarios, relative yield losses for winter wheat12 12 are less 

13 than one percent. For soybeans, yield losses for these scenarios range from just above 1 percent 

14 to below one percent (U.S. EPA 2014, section 6.5). 

15 The WREA estimates of O3-attributable percent yield loss based on average W126 values 

16 estimated for just meeting the current standard are relatively small (0.0 – 2.72%, U.S. EPA 2014, 

17 section 6.5, Appendix 6B). In considering these estimates, we recognize the significant 

18 uncertainties associated with several aspects of the analyses. Because the W126 estimates 

19 generated in the air quality analyses are inputs to the vegetation risk analyses for biomass loss, 

20 crop loss and foliar injury, any uncertainties in the air quality analyses are propagated into the 

21 those analyses (U.S. EPA 2014, Table 6-27, section 8.5). 

22 • To what extent are the exposures and risks remaining upon simulating just 

23 meeting the current O3 standard important from a public welfare perspective?

24 From a public welfare prospective, the O3 attributable risks to crops estimated for 

25 conditions that just meet the current standard are small. As discussed in the WREA and 

26 summarized above, there are multiple areas of uncertainty associated with these estimates, 

27 including those associated with the model-based adjustment methodology as well as those 

28 associated with projection of yield loss at the estimated O3 concentrations (U.S. EPA, 2014, 

29 Table 6-27, section 8.5). 

 

12 Among the major crops, because winter wheat and soybeans are more sensitive to ambient O3 levels than 

other crops we include these crops for this discussion. 
5-35 

1 • What are the ecosystem services potentially affected by O3-related crop yield loss 

2 and to what extent are they important from a public welfare perspective? To 

3 what degree can the magnitude of the O3 effect on these services be qualitatively 

4 or quantitatively characterized?

5 The WREA presents national-scale estimates of the effects of biomass loss on timber 

6 production and agricultural harvesting, as well as on carbon sequestration (U.S. EPA 2014, 

7 section 6.5). Because the forestry and agriculture sectors are related, and trade-offs occur

8 between the sectors, the WREA also calculated the resulting market-based welfare effects of O3

9 exposure in the forestry and agriculture sectors. Overall effect on agricultural yields and 

10 producer and consumer surplus depends on the (1) ability of producers/farmers to substitute 

11 other crops that are less O3 sensitive, and (2) responsiveness, or elasticity, of demand and supply 

12 (U.S. EPA, 2014, sections 6.5, 8.2.1.3). Estimated O3-attributable economic welfare impacts on 

13 agricultural sectors associated with air quality conditions model-adjusted to just meet the existing 

14 and potential alternative W126 standard levels were largely similar between the forestry and 

15 agricultural sectors. Estimates of consumer surplus, or consumer gains, were generally higher 

16 under those conditions (compared to recent conditions) in both sectors because higher 

17 productivity under lower O3 concentrations increased total yields and reduced market prices

18 (U.S. EPA 2014, Table 6-16). Because demand for most forestry and agricultural commodities 

19 is not highly responsive to changes in price, there were more examples for which producer 

20 surplus, or producer gains, decline. For agricultural welfare, annualized combined consumer and 

21 producer surplus gains were estimated to be $2.6 trillion for model adjustment to meet the 

22 current standard. Combined gains were essentially unchanged in comparisons of the current 

23 standard scenario to the average W126 scenario for 15 ppm-hrs, but gains increased by $21 

24 million for the W126 scenario for 11 ppm-hrs and $231 million for the W126 scenario for 7 

25 ppm-hrs. In some cases, lower prices reduce producer gains more than can be offset by higher 

26 yields (U.S. EPA, 2014, Table 6-17). 

27 The WREA discusses multiple areas of uncertainty associated with these estimates (also 

28 summarized above), including those associated with the model-based adjustment methodology as 

29 well as those associated with projection of yield loss at the estimated O3 concentrations (U.S. 

30 EPA, 2014, Table 6-27, section 8.5). 

31 5.4 VISIBLE FOLIAR INJURY 

32 Visible foliar injury resulting from exposure to O3 has been well characterized and 

33 documented over several decades of research on many tree, shrub, herbaceous, and crop species 

34 (U.S. EPA, 2013, 2006, 1996, 1984, 1978). The significance of O3 injury at the leaf and whole 

35 plant levels depends on an array of factors and there is difficulty in relating visible foliar injury 
5-36 

1 symptoms to other vegetation effects such as individual tree growth, or effects at population or 

2 ecosystem characteristics (U.S. EPA, 2013, p. 9-39). Visible foliar injury by itself, however, can 

3 impact the public welfare through damaging or impairing the intended use of the affected entity 

4 or the service it provides. For example, ways by which O3-induced visible foliar injury may 

5 impact the public welfare thus include: 1) visible damage to ornamental species or leafy crops 

6 (spinach, lettuce, tobacco) that affects the economic value, yield, or usability of that plant (U.S. 

7 EPA 2007, section 7.4.1; Abt Associates, Inc. 1995); 2) visible damage to plants with special 

8 cultural significance (e.g., those used in tribal practices); 3) visible damage to species occurring 

9 in natural settings valued for their scenic beauty and/or recreational appeal, including in areas 

10 specially designated for more protection (i.e., federal Class I areas) (73 FR 16490). Given 

limitations in the available information pertaining to the first two categories, 13 11 the discussions of 

12 the evidence and exposure/risk information in sections 5.4.1 and 5.4.2 below focus primarily on 

13 what is known about visible foliar injury that has been shown to occur in natural settings valued 

14 for their scenic beauty and/or recreational appeal.

15 At the time of the last review, the following was known:

16 1) Ozone causes diagnostic visible injury symptoms on studied bioindicator species. 

17 2) Soil moisture is a major confounding effect that can decrease the incidence and 

18 severity of visible foliar injury under dry conditions and visa versa. 

19 3) The most extensive dataset regarding visible foliar injury incidence across the U.S. 

20 was that collected by the U.S. Forest Service (USFS) Forest Health Monitoring/Forest 

21 Inventory and Analysis (FHM/FIA) Program. 

22 4) Staff analyses of county level air quality data and USFS biomonitoring data showed 

23 that for each year within a four year period (2001 – 2004) the percentage of counties 

24 having a biosite with visible foliar injury ranged between 11-30% at an 8-hour average 

25 annual level of 0.074 ppm (U.S. EPA, 2007, section 7.6.3.2). 

26 In the remainder of this section, we consider how the currently available evidence and 

27 exposure/risk information informs our understanding of the relationship that exists between 

28 visible foliar injury and exposures to O3 in ambient air and consideration of the adequacy of 

29 protection provided by the current standard. The policy-relevant evidence and weight-of30 evidence conclusions drawn from the ISA are discussed in section 5.4.1, and the exposure/risk 

31 and associated ecosystem services estimates from the second draft WREA, are discussed in 

32 section 5.4.2. Important uncertainties and limitations in each type of available information are 

33 also discussed in these two sections. 

 

13 Qualitative information regarding potential cultural impacts of O3-induced visible foliar injury is noted in 

section 5.5 and Appendix 5A). 
5-37 

1 5.4.1 Evidence-based Considerations 

2 • To what extent has scientific information become available that alters or 

3 substantiates our previous conclusions of O3-related visible foliar injury and of 

4 factors that influence associations between O3 exposures or concentrations and 

5 visible foliar injury? 

6 Recent research continues to build and substantiate the previous conclusions and findings 

7 drawn from several decades of research on many tree, shrub, herbaceous, and crop species (U.S. 

8 EPA, 2013, 2006, 1996, 1984, 1978) that O3-induced visible foliar injury symptoms are well 

9 characterized and considered diagnostic on certain bioindicator plant species. Diagnostic usage 

10 for these plants has been verified experimentally in exposure-response studies, using exposure 

11 methodologies such as continuous stirred tank reactors (CSTRs), open-top chambers (OTCs), 

12 and free-air fumigation (FACE). Although there remains a lack of robust exposure-response 

13 functions that would allow prediction of visible foliar injury severity and incidence under 

14 varying air quality and environmental conditions, experimental and observational evidence has 

15 clearly established a consistent association of the presence of visible injury symptoms with O3

16 exposure, with greater exposure often resulting in greater and more prevalent injury (U.S. EPA 

17 2013, section 9.4.2). This new research includes: 1) controlled exposure studies conducted to 

18 test and verify the O3 sensitivity and response of potential new bioindicator plant species; 2) 

19 multi-year field surveys in several National Wildlife Refuges documenting the presence of foliar 

20 injury in valued areas; 3) ongoing data collection and assessment by the USDA Forest Service 

21 FHM/FIA program, including multi-year trend analysis (U.S. EPA 2013, section 9.4.2). These 

22 recent studies, in combination with the entire body of available evidence, thus form the basis for 

23 the ISA determinations of a causal relationship between ambient O3 exposure and the occurrence 

24 of O3-induced visible foliar injury on sensitive vegetation across the U.S. (U.S. EPA 2013, p. 9-

42). 25 

26 With regard to evidence from controlled exposure studies, a recent study of 28 plant 

27 species confirmed prior findings of O3 causing predictable diagnostic visible foliar injury 

28 symptoms on some species of plants. This study selected 28 plant species, most of which grow 

29 naturally throughout the northeast and midwest US, including in national parks and wilderness 

30 areas, that were suspected of being O3 sensitive, and exposed them to four different O3

31 concentrations (30, 60, 90, and 120 ppb) in continuously stirred tank reactor (CSTR) chambers 

32 (Kline et al., 2008). Two experiments were conducted in each year of the study (2003 and 2004). 

33 Plants were exposed to O3 for 7 hours a day, five days a week over the course of each 

34 experiment. Specifically, in 2003, the first experiment lasted from July 14 to August 21and 

35 included 29 days of O3 exposure and the second from September 9 to 30 and included 16 

36 exposure days. In 2004, the first experiment was conducted from July 13 to August 10 with 21 
5-38 

1 O3 exposure days and the second from August 27 to September 24, including 21 days of O3

2 exposure. Though the exposures were cumulative over the course of the study, exposures were 

3 reported only in terms of the target exposure concentration for each experiment. The study 

4 reported O3-induced responses in 12, 20, 28 and 28 of the 28 tested species at the 30, 60, 90 and 

120 ppb exposure concentrations14 5 , respectively. Based on their findings, the authors suggest 

6 that American sycamore, aromatic sumac, bee-balm, buttonbush, common milkweed, European 

7 dwarf elderberry, New England aster, snowberry and swamp milkweed would make the most 

8 useful bioindicator species. Some of these species are native to U.S. locations designated of 

9 national significance (discussed further below). The staff additionally concludes that given that 

10 the exposure protocol was designed to create a continuous exposure level, not a fluctuating one, 

11 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.15 12 

13 While this type of controlled study provides clear evidence of cause and effect, it also has 

14 limitations. The authors, recognizing this cautioned that “extrapolation of these CSTR results to 

15 the field must be done carefully, since CSTR/greenhouse conditions ... are not representative of 

16 natural environmental conditions” (Kline et al., 2008). 

17 A string of recently published multi-year field studies provide a complimentary line of 

18 field-based evidence by documenting the incidence of visible foliar injury symptoms on a variety 

19 of O3-sensitive species over multiple years and across a range of cumulative, seasonal exposure 

20 values in several eastern and midwestern Class I national wildlife refuges (NWRs) (U.S. EPA 

21 2013, section 9.4.2.1; Davis and Orendovinci 2006; Davis 2007a, b; Davis 2009). Some of these 

22 studies also included information regarding soil moisture stress using the Palmer Drought 

23 Severity Index (PDSI). While environmental conditions and species varied across the four 

24 NWRs, visible foliar injury was documented to a greater or lesser degree at each site. As 

25 discussed further below, visible foliar injury incidence in these types of areas has greater 

26 significance to the public welfare. 

27 • To what extent have important uncertainties identified in the last review been 

28 reduced and/or new uncertainties emerged?

29 The studies mentioned above also provide additional information regarding an important 

30 uncertainty indentified in the previous review, i.e., the role of soil moisture in influencing visible 

 

14 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. 

15 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. 
5-39 

1 foliar injury response (U.S. EPA 2013, section 9.4.2). These studies confirm that adequate soil 

2 moisture creates an environment conducive to greater visible foliar injury in the presence of O3

3 than drier conditions. As stated in the ISA, “[a] major modifying factor for O3-induced visible 

4 foliar injury is the amount of soil moisture available to a plant during the year that the visible 

5 foliar injury is being assessed ... because lack of soil moisture generally decreases stomatal 

6 conductance of plants and, therefore, limits the amount of O3 entering the leaf that can cause 

7 injury” (U.S. EPA, 2013, p. 9-39). As a result, “many studies have shown that dry periods in 

8 local areas tend to decrease the incidence and severity of O3-induced visible foliar injury; 

9 therefore, the incidence of visible foliar injury is not always higher in years and areas with higher 

10 O3, especially with co-occurring drought (Smith, 2012; Smith et al., 2003)” (U.S. EPA, 2013, p. 

11 9-39). This “...partial ‘protection’ against the effects of O3 afforded by drought has been 

12 observed in field experiments (Low et al., 2006) and modeled in computer simulations 

13 (Broadmeadow and Jackson, 2000)” (U.S. EPA, 2013, p. 9-87). In considering the extent of any 

14 protective role of drought conditions, however, the ISA also notes that other studies have shown 

15 that “drought may exacerbate the effects of O3 on plants (Pollastrini et al., 2010; Grulke et al., 

16 2003)” and that “[t]here is also some evidence that O3 can predispose plants to drought stress 

17 (Maier-Maercker, 1998)”. Accordingly, the ISA concludes that “the nature of the response is 

18 largely species-specific and will depend to some extent upon the sequence in which the stressors 

19 occur” (U.S. EPA, 2013, p. 9-87). Such uncertainties associated with describing the potential for 

20 foliar injury and its severity or extent of occurrence for any given air quality scenario due to 

21 confounding by soil moisture levels make it difficult to identify an appropriate degree of annual 

22 protection (as well as ambient O3 exposure conditions that might be expected to provide that 

23 protection). 

24 • To what extent does the available evidence indicate the occurrence of O3-related 

25 visible foliar injury attributable to cumulative exposures at lower ambient O3

26 concentrations than previously established or to exposures at or below the level of 

27 the current standard?

28 Recently available evidence confirms that available in previous reviews that visible foliar 

29 injury can occur when sensitive plants are exposed to elevated O3 concentrations in a 

30 predisposing environment (i.e. adequate soil moisture (U.S. EPA, 2013, section 9.4.2). Recent 

31 evidence also continues to indicate the occurrence of visible foliar injury at cumulative ambient 

32 O3 exposures previously established. Since the 2006 O3 AQCD, results from several multi-year 

33 field surveys and experimental screenings of O3-induced visible foliar injury on vegetation also 

34 show that visible foliar injury can occur under conditions where the annual 8-hour average O3

35 concentrations are at or below the level of the current standard, as discussed here. Limited 
5-40 

1 information exists regarding the incidence of visible foliar injury occurring in areas that have 

2 design values that meet the current 3-year average 8-hour standard. 

3 To facilitate comparison with other studies reporting foliar injury response to W126 

4 cumulative exposures, we obtained air quality data from EPA’s AQS database for monitors in 

each study location and calculated the 12-hr W126 values and obtained the maximum 4th 5 high 8-

6 hour average values for a subset of the most recent years included in each study (Table 5-6). As 

7 the shaded rows in Table 5-6 below show, in the years 2002/2003 and 2004 in the Cape Romain 

8 National Wildlife Refuge in SC, and the Seney National Wildlife Refuge in Michigan, 

respectively, the 4th 9 highest daily maximum 8-hour average O3 concentrations were at or below 

10 the level of the current standard. We additionally note that the Cape Romain site met the current 

11 standard of 75 in every 3-year period during the study and has consistently met the standard from 

2001 to 2012.16 12 Under these air quality conditions, three species exhibited O3-induced stipple 

13 (winged sumac, Chinese tallow tree, and wild grape). In 2002, 32% of the examined wild grape 

14 plants, 20% of the winged sumac plants, and 4.6% of the Chinese tallow tree plants, respectively, 

15 were symptomatic (Davis 2009). At the same time, the 12 hour W126 value was 20 ppm-hrs. In 

16 2003, when air quality was somewhat improved, foliar injury declined, with only 13.3% of wild 

grape showing ozone stipple at a maximum 4th 17 high 8-hr of 74 ppb and a W126 of 11 ppm-hr. 

18 The PSDI values were 0.27 and 2.45 in 2002 and 2003, respectively. These values show that 

19 2003 was a wetter year than 2002, though 2002 would have been considered within the normal 

20 soil moisture range. 

21 At the Seney NWR site, by comparison, the annual W126 level was similar in 2004 to 

22 that at Cape Romain in 2003, and the annual 8-hour average level was below that of the current 

23 standard, though the 3-year average design values were above that of the current standard for that 

24 year. Not surprisingly, given the lower O3 air quality in 2004, the Seney study reported injury 

25 ranging from about 2% on common milkweed to about 6% on spreading dogbane. Though this 

26 study does not provide the PDSI values, the authors provided some discussion of a possible 

27 relationship stating that “the incidence of ozone injury on spreading dogbane, but not other 

28 species, was weakly, but not significantly, related to the drought index (PDSI)....However this 

29 relationship was too weak to be used for predictive purposes.” The authors then conclude that 

30 “[n]evertheless, the threshold SUM06 ozone level needed to induce stipple on sensitive plants 

31 within the Seney refuge is likely 5000 ppb-hrs under the environmental conditions of these 

32 surveys.” On the basis of the above, the staff concludes that these studies confirm that visible 

33 foliar injury has been shown to occur in the field at W126 index values ranging down to 10 ppm-

 

16Design 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
5-41 

1 hrs and provide limited evidence that such foliar injury can occur in areas with special public 

2 welfare significance during periods that meet the current standard. 

3 Table 5-6. Visible foliar injury incidence in four National Wildlife Refuges.

Name/ Site #/ Ref.A Year B 4th highest daily maximum 

8-hr average

12 hr. W126 % Plants with 

visible injury

Cape Romain NWR, SC/ 

450190046 (Davis, 2009)

2002 0.075 ppm 20 ppm-hr 5 - 32

2003 0.074 ppm 11 ppm-hr 3-13

Moosehorn NWR, ME/ 230090102 

(Davis, 2007a)

2002 0.1 ppm 24 ppm-hr 0 - 17

2003 0.083 ppm 22 ppm-hr 0 - 13

2004 0.082 ppm 14 ppm-hr 3 - 10

Seney NWR, MI/ 261530001 

(Davis, 2007b)

2002 0.083 ppm 11 ppm-hr 0 - 13

2003 0.076 ppm 15 ppm-hr 1 - 6

2004 0.074 ppm 10 ppm-hr 2 - 6

NWR, NJ/ 340010005/ 

(Davis and Orendovici, 2006)

2001 0.095 ppm 39 ppm-hr 0 – 45

2002 0.092 ppm 53 ppm-hr 0 – 4

2003 0.085 ppm 36 ppm-hr 0 - 4

AStudies (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-hr and W126 forms : http://www.epa.gov/ttn/airs/airsaqs/ BOnly recent years with available W126 data were included in Table.

4 

5 By far the most extensive field-based dataset of visible foliar injury incidence is that 

6 obtained by of the USDA Forest Service FHM/FIA biomonitoring network program. A trend 

7 analysis of data from the sites located in the Northeast and North Central U.S. for the 16 year 

8 period (1994-2009) (Smith, 2012) provides additional evidence of foliar injury occurrence in the 

9 field as well as some insight into the influence of changes in air quality and soil moisture on 

10 visible foliar injury and the difficulty inherent in predicting foliar injury response under different 

11 air quality/soil moisture scenarios (Smith, 2012; U.S. EPA 2013, section 9.2.4.1). In this study 

12 ambient exposures were expressed in terms of the SUM06 cumulative index coupled with a 

13 measure of the number of peak hourly concentrations above 100 ppb (N100). Soil moisture 

14 conditions were generated using both the Palmer Drought Severity Index (PSDI) and the plant 

moisture availability index (MI). Foliar injury was expressed in terms of the biosite index (BI)17 15 . 

 

17 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.
5-42 

1 The authors observed that over this 16-year period, “injury indices have fluctuated annually in 

2 response to seasonal ozone concentrations and site moisture conditions. Sites with and without 

3 injury occur at all ozone exposures but when ambient concentrations are relatively low, the 

4 percentage of uninjured sites is much greater than the percentage of injured sites; and regardless 

5 of ozone exposure, when drought conditions prevail, the percentage of uninjured sites is much 

6 greater than the percentage of injured sites” (Smith, 2012). The authors further note that while 

7 “both site moisture and ozone exposure play a role in foliar injury expression ... the interplay 

8 among these three factors is unique for each year and possibly each site. Extreme moisture 

9 deficits decrease foliar injury, ... [and] ... [i]n no year do high ozone exposures override the 

10 controlling effect of site moisture, although at the other end of the scale, injury severity is 

11 minimized under conditions of low ozone exposure regardless of site moisture conditions. This 

12 implies a necessary threshold of ozone exposure for injury to occur....” “In a similar analysis, 

13 Rose and Coulston (2009) reported a high percentage of biosites with injury across the Southern 

14 region in 2003, a year when SUM06 values >10 ppm-h were widespread at the same time that 

15 the land area was in moisture surplus or balance.” Thus, Rose and Coulston (2009) also “found 

16 evidence that it is the co-occurrence of sufficient moisture and elevated ozone that determine 

17 whether injury occurs to bioindicator plants, not ozone exposure alone.” Regarding the role of 

18 peak ozone concentrations (>100 ppb O3), Smith (2012) reported that over the 16-year period 

19 concentrations above 100 ppb have declined, and that this “... may account for the observed 

20 decrease in the severity of ozone-induced foliar injury to ozone sensitive bioindicator plants in 

21 eastern forests.” They also note that “[t]here is no compelling evidence, however, that moderate 

22 ozone concentrations, as reflected in seasonal mean SUM06 data, are on the decline....” “This 

23 may explain why injury continues to be detected on many of the same sites every year...” The 

24 authors thus conclude that, “[a]lthough it is reasonable to remain concerned about long-term 

25 impacts of ozone pollution on our forest ecosystems, the findings of this biomonitoring survey 

26 point to a declining risk of probable impact on eastern forests over the 16-year period from 1994 

27 to 2009.” 

28 In a similar assessment of the USDA Forest Service FHM/FIA data in the West, six years 

29 (2000 to 2005) of biomonitoring data for O3 injury were evaluated for the three coastal states of 

30 California, Oregon and Washington (Campbell et al., 2007; U.S. EPA 2013, section 9.4.2.1). 

31 Campbell et al., 2007 found that “...ozone injury occurs frequently (25 to 37 percent of sampled 

32 biosites) in California forested ecosystems demonstrating that ozone is present at phytotoxic 

33 levels.” This study concluded that, “in California, an estimated 1.3 million acres of forest land 

34 and 596 million cubic feet of wood are at moderate to high risk to impacts from ozone. 

35 However, [m]ore years of data are needed to discern any trends.” Though this study does not 

36 discuss the role of soil moisture in describing the results, the criteria used to select the 
5-43 

1 biomonitoring sites include one that considers soil conditions. The best sites are identified as 

2 those with low drought potential and good fertility. Thus, given the relatively high O3

3 concentrations that occur in California and the likelihood that many of the biomonitoring sites 

4 occur in areas that have sufficient soil moisture, the high percentage of sampled biosites with 

foliar injury is not unexpected.18 5 

6 These recent studies continue to provide evidence of O3-induced foliar injury occurring in 

7 many areas across the U.S. and augment our understanding of O3-related visible foliar injury and 

8 of factors that influence associations between O3 exposures or concentrations and visible foliar 

9 injury such as soil moisture. 

10 • To what extent does currently available evidence suggest locations where the 

11 vulnerability of sensitive species, ecosystems and/or their associated services to 

12 O3-related visible foliar injury would have special significance to the public 

13 welfare?

14 As mentioned above, federally designated Class I areas are afforded stringent protections 

15 under the 1977 amendments to the Clean Air Act (Act). The Act gives federal land managers of 

16 Class I areas “the responsibility to protect all air quality related values (AQRVs)...from 

17 deterioration.... In order to determine if deterioration is occurring, baseline AQRVs must be 

18 established” (Davis, 2009). Because of this need and the significance of these areas, studies 

19 often focus on these sites. For example, a study by Kohut (2007) was undertaken to assess the 

20 risks of O3-induced visible foliar injury on O3 bioindicators (i.e., O3-sensitive vegetation) in 244 

21 parks managed by the NPS. Kohut (2007) estimated O3 exposure using hourly O3 monitoring 

22 data collected at 35 parks from 1995 to 1999, estimated O3 exposure at 209 additional parks 

23 using kriging, a spatial interpolation technique, and qualitatively assessed risk. Kohut (2007) 

24 applied a subjective evaluation based on three criteria: (1) the frequency of exceedance of foliar 

injury thresholds19,20 25 using several O3 exposure metrics (i.e., SUM06, W126 and N100), (2) the 

26 extent that low soil moisture constrains O3 uptake during periods of high exposure, and (3) the 

27 presence of O3 sensitive species within each park. Based on these criteria, Kohut (2007) 

 

18 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).

19 Kohut (2007) uses the term “foliar injury thresholds”. In the WREA foliar injury analyses, we use the 

term “benchmarks” or “benchmark criteria” to avoid implying that foliar injury could not occur below these levels. 

20 Consistent with advice from CASAC (Frey and Samet, 2012a), the WREA modified the approach used 

by Kohut (2007) to apply the W126 metric. The WREA analysis, described in section 5.2.2 below developed and 

considered these different W126 benchmarks for foliar injury 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 (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.
5-44 

1 concluded that the risk of visible foliar injury was high in 65 parks (27 percent), moderate in 46 

2 parks (19 percent), and low in 131 parks (54 percent). Thus, while this study suggests that there 

3 may be a reason for concern in as much as 46% of the parks, there were a number of important 

4 limitations associated with this study (described in footnotes 8 and 9 below) that weakened this 

5 conclusion. Given the importance of this kind of assessment, the WREA used Kohut (2007) as 

6 the conceptual basis for the subsequent WREA national-scale and screening-level analyses, 

7 though numerous modifications were made to the approach to make it applicable to the context 

8 of this O3 NAAQS review (see section 5.4.2 below). 

9 In addition, as described above, several recently published studies (U.S. EPA 2013, 

10 section 9.4.2.1; Davis and Orendovici 2006; Davis 2007a, b; Davis 2009, Kohut 2007) were 

11 conducted in federally protected areas including federally designated Class I areas and national 

12 parks. These studies confirm that visible foliar injury has been observed in these areas under 

13 annual air quality conditions with ambient concentrations at or below the level of the current 

14 standard and at W126 index values within the CASAC range recommended in past reviews. 

15 This evidence continues to suggest that O3-sensitive species and their associated ecosystems and 

16 services continue to remain vulnerable to visible foliar injury incidence in areas that have been 

17 afforded special protection by Congress and that have special significance to the public welfare. 

18 5.4.2 Exposure-/Risk-based Considerations

19 The WREA presents a number of analyses considering air quality conditions associated 

20 with increased prevalence of visible foliar injury and potential associated welfare impacts (see 

21 Table 5-7 below, U.S. EPA, 2014, Chapter 7). An initial analysis included the development of 

22 benchmark criteria reflecting different prevalences of visible foliar injury or of the occurrence of 

23 elevated injury in conjunction with different W126 exposures and in some cases, soil moisture 

24 conditions. These criteria were then used in a national scale screening level assessment to 

25 characterize potential risk of foliar injury incidence under 2006-2010 conditions in 214 national 

26 parks. The last analysis was a case study assessment on three national parks, which also 

27 provides limited characterization of the associated ecosystem services. Despite the limitations 

28 and uncertainties associated with these analyses, and recognizing that the air quality conditions 

29 in most cases (prior to any model adjustment) did not meet the current standard, staff believes 

30 that they help inform our understanding of the relationship between soil moisture and foliar 

31 injury incidence, as well as provide limited support for our conclusion of a risk of visible foliar 

32 injury incidence under air quality conditions likely to meet the current standard in areas of 

33 special significance to the public welfare.
5-45 

1 Table 5-7. Exposure, risk and ecosystem services analyses related to visible foliar injury. 

Ecosystem Level Effects Ecosystem Services

WREA 

estimates

Proportion of USFS biosites with 

foliar injury incidence, degree of 

injury, and soil moisture at various

W126-based benchmarks

Percent of 214 national parks 

exceeding criteria based on USFS 

biosite dataset analysis during years 

considered

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.

B 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. 

2 

3 • For what air quality scenarios were exposures and risks estimated? What

4 approaches were used to estimate W126 exposures for those conditions? What 

5 are associated limitations and uncertainties?

6 Three types of foliar injury analyses were performed in the WREA and are considered 

7 below. They include an analysis using U.S. Forest Service biosite data, a National Park 

8 screening-level assessment and a National Park case study (focused on three parks). The 

9 analysis of U.S. Forest Service (USFS) biosite data was done using O3 concentrations estimated 

10 for a national-scale surface of concentrations (at a 12 km grid cell resolution in contiguous U.S.) 

11 using interpolation methodology applied to concentrations at O3 monitor locations (U.S. EPA, 

12 2014, section 4.3.2, Appendix 4A). The analysis of USFS data used surfaces for each year from 

13 2006 through 2010 (U.S. EPA, 2014, Appendix 4A, section 4.2). In both the National Park 

14 screening-level assessment and the case study analysis, observations related to air quality were 

made for five air quality scenarios by the methodology summarized in Table 5-4 above.21 15 

16 The W126 index values in the individual years from 2006 to 2010 ranged from less than 

17 5 ppm-hrs to 25 ppm-hrs. Concentration estimates varied appreciably across the five years with 

 

21 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 

model-adjusted concentrations for the 4 other scenarios. The development of model adjusted concentrations was 

done for each of 9 regions independently (see U.S. EPA, 2014, 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 

Voronoi Neighbor Averaging (VNA) spatial interpolation technique (details on the VNA technique are presented in 

U.S. EPA, 2014 Appendix 4A). 
5-46 

1 the median index values across grid cells ranging from a low of 5.5 ppm-hrs in 2009 to 11 ppm2 hrs in 2006 (U.S. EPA, 2014, Appendix 4A, section 4.2). During the recent conditions period 

3 (2006 through 2008), the average W126 index values (across the three-year recent conditions 

4 period) at the monitor locations ranged from below 10 ppm-hrs to 48.6 ppm-hrs (U.S. EPA 2014, 

5 Figure 4-4 and Table 4-1). After model-adjustment of the 2006-2008 data to just meet the 

6 current standard in each region, and subsequent application of the VNA technique to the current 

7 standard scenario monitor location values, the average W126 values were below 15 ppm-hrs 

8 across the national surface with the exception of a very small area of the southwest region (near 

9 Phoenix) where the average W126 values was near or just above 15 ppm-hrs. A lowering of the 

10 highest values occurred with application of the interpolation method as a result of estimating 

W126 values at a 12x12 km2 11 grid resolution rather that at the exact location of a monitor. This 

12 indicates one uncertainty associated with this aspect of the approach to estimating W126 values 

13 for the model-adjusted air quality just meeting the current standard. Other uncertainties are 

14 summarized in section 5.2.2 above. 

15 • What are the nature and magnitude of the cumulative exposure- and risk-related 

16 estimates for visible foliar injury under recent conditions or conditions meeting 

17 the current O3 standard? 

18 As an initial matter, we consider the analysis of the biomonitoring site data from the 

USDA Forest Service FHM/FIA Network, described in section 7.2 of the WREA.22 19 Using this 

20 dataset and associated data for soil moisture conditions during the sample years along with 

21 ambient air O3 concentrations based on monitoring data from 2006 to 2010 and spatial 

22 interpolation methodology (as described above), the proportion of sites with any or elevated 

23 foliar injury are observed to increase with increasing annual W126 index values up to specific 

24 values after which there is little change in proportion of affected sites with higher W126 values 

25 (see Figure 5-5 below; U.S. EPA, 2014, section 7.2). The proportion of sites metric is derived by 

26 first ordering the data (across sites and sample years) by W126 index value estimated for that site 

27 and year. Then for each W126 index value the proportion of sites exceeding the selected biosite 

28 index value for all observations at or below that W126 index value is calculated. The WREA

29 repeated this using a biosite index value greater than zero, indicating presence of any foliar 

30 injury, and an index value >5, corresponding to a USFS cutoff for “low” or more injury (USFS, 

31 2011). 

32 When looking only at presence or absence of foliar injury (“any injury”) with the 

33 exception of 2008, the proportion of sites across all W126 index values with foliar injury exceeds 

 

22 Data were not available for several western states (Montana, Idaho, Wyoming, Nevada, Utah, Colorado, 

Arizona, New Mexico, Oklahoma, and portions of Texas).
5-47 

1 15 percent; in 2006, it exceeds 20 percent, while in 2008 the proportion of sites with foliar injury 

2 across all W126 index values was just below 15 percent (U.S.EPA, 2014, section 7.2.3, Figure 7-

3 8). When categorized by moisture levels, the data demonstrate a distinct pattern. In general, the 

4 WREA concludes that the results of these foliar injury analyses demonstrate a similar pattern –

5 the proportion of sites showing the presence of any foliar injury (biosite>0) or at least little foliar 

6 injury (biosite>5) increases from zero to about 20% and 6% respectively (Figure 5-5 below). 

7 This increase occurs with increasing W126 index values up to approximately10 ppm-hrs for any

8 foliar injury (biosite index >0), with little change in proportion of sites with any injury at higher

9 W126 values. The data for sites during normal moisture years are very similar to the dataset as a 

10 whole, with an overall proportion of close to 18 percent for presence of any foliar injury, and 

11 close to 6 percent for sites exceeding a biosite index of 5. Among the sites with relatively wet 

12 season (average Palmer Z => 1), the highest proportion of sites observed is much higher for both 

13 index categories of injury and the relationship with annual W126 index value is much steeper. 

14 Much lower proportions of sites are reached for both injury categories at sites with relatively dry 

15 seasons (average Palmer Z < -1.24), potentially indicating that drought may provide some 

16 protection from foliar injury as discussed in the ISA (U.S. EPA, 2014, section 7.2.3, Figures 7-10 

17 and 7-11). This information provides insight into the relationship between soil moisture and 

18 foliar injury and the issue of whether drought provides protection from foliar injury. For both 

19 categories of injury, there is relatively little change in the proportion of sites beyond a W126 of 

20 20 ppm-hrs. There are two important observations that can be made from these analyses: (1) the 

21 proportion of sites exhibiting foliar injury rises rapidly at increasing W126 index values below 

22 10 ppm-hrs, and (2) there is relatively little change in the proportions above W126 index values 

23 of 20 ppm-hrs.
5-48 

1 

2 

3 Figure 5-5. Cumulative proportion of sites with a) any foliar injury or b) elevated injury 

4 present, by moisture category (U.S. EPA 2014, Figures 7-10 and 7-11).

5 We additionally consider the WREA screening-level assessment in 214 parks in the 

6 contiguous U.S. that employed benchmark criteria developed from the above analysis (Table 5-

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

0 10 20 30 40

0.00 0.02 0.04 0.06 0.08

Biosite Index > 5

W126 (ppm-hrs)

Proportion of sites

All

Wet

Normal

Dry
5-49 

8).23,

24 1 For example, annual O3 concentrations corresponding to a W126 seasonal index value of 

2 10.46 ppm-hrs represents the O3 exposure concentration above which a consistent percentage of 

3 all biosites (17.7 percent) showed any injury, without consideration of soil moisture. The W126 

4 benchmarks across the six scenarios range from 3.05 ppm-hrs (five percent of biosites, normal 

5 moisture, any injury) up to 46.87 ppm-hrs (five percent of biosites, dry, elevated injury). For the 

6 scenario of 10% biosites with injury, W126 values were approximately 4, 6, and 25 ppm-hours 

7 for wet, normal and dry years, respectively. The national-scale screening-level assessment 

8 includes 42 parks with O3 monitors and 214 parks with O3 exposure estimated from the 

9 interpolated O3 surface for individual years from 2006 to 2010 (U.S. EPA, 2014, Appendix 7A). 

10 These data were combined with lists from the NPS of the parks containing O3 sensitive 

11 vegetation species (NPS, 2003, 2006). Based on NPS lists, 95 percent of the parks in this 

12 assessment contain at least one O3-sensitive species. This analysis for recent air quality 

13 conditions, estimates that 58 percent of parks exceeded the benchmark criteria corresponding to 

14 the base scenario (W126>10.46 ppm-hrs, 17.7 percent of biosites, all moisture categories, any 

15 injury) for at least three years in the period from 2006 to 2010, and 34 percent of parks would 

16 exceed the benchmark criteria for the elevated injury scenario (five percent of biosites, multiple 

17 moisture categories, elevated foliar injury) for at least three years (U.S. EPA, 2014, section 

7.3.2).25 18 Based on model-adjustments to meet the current standard, none of the 214 parks have 

19 average W126 index values that would exceed the annual benchmark criteria for the base 

20 scenario (W126 >10.46 ppm-hrs).

21 

 

23 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.

24 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 (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.

25 The lack of national surfaces of O3 concentrations for each of three years in each scenario (as described 

earlier and summarized in Table 5-3) precluded derivation of similar estimates (of parks with W126 index values 

above any of the benchmark criteriaafter model-adjustment of air quality to just meet the current standard, although 

we note that the W126 averaged across three years in all 214 parks was below 10.46 ppm-hrs (the individual year 

benchmark criteria for the base scenario) (U.S. EPA, 2014, section 7.3.2). 
5-50 

1 Table 5-8. Benchmark criteria for O3 exposure and relative soil moisture used in 

2 screening-level assessment of parks (from U.S. EPA 2014, Table 7-5).

Description Normal

(-1.25<Palmer Z<1)

Wet 

(Palmer Z>1)

Dry

(Palmer Z<-1)

Any injury, without 

considering soil 

moisture conditions 

(Base Scenario)

17.7% of all biosites with any foliar 

injury (consistent proportion of all 

biosites showed injury in years with 

W126 index value above this value)

W126>10.46

(soil moisture not considered)

Any injury, 

considering soil 

moisture conditions

5% of biosites showing any foliar 

injury, by relative moisture category

W126>3.05 W126>3.76 W126>6.16

10% of biosites showing any foliar 

injury, by relative moisture category

W126>5.94 W126>4.42 W126>24.61

15% of biosites showing any foliar 

injury, by relative moisture category

W126>8.18 W126>4.69 N/A

20% of biosites showing any foliar 

injury, by relative moisture category

N/A W126>5.65 N/A

5% of biosites, 

Injury ≥ 5

5% of biosites showing foliar injury 

equal or greater than 5 on the biosite 

injury index (e.g., 5% of leaf shows 

injury in 10% of the leaves), by 

relative moisture category

W126>12.23 W126>7.02 W126>46.87

3 

4 Lastly, we consider the WREA case study analysis which focused on characterizing the 

5 ecosystem services potentially associated with visible foliar injury in three specific national 

6 parks (case study assessment). The parks included were Great Smoky Mountains National Park 

7 (GRSM), Rocky Mountain National Park (ROMO), and Sequoia/Kings National Parks (SEKI). 

8 For each park, the potential impact of O3-related foliar injury on recreation (cultural services) 

9 was considered in light of information on visitation patterns, recreational activities and visitor 

10 expenditures. For example, visitor spending in 2011 exceeded $800 million, $170 million and 

11 $97 million dollars in GRSM, ROMO and SEKI, respectively. This assessment also included 

12 percent cover of species sensitive to foliar injury and focused on the overlap between recreation

13 areas within the park and elevated W126 concentrations. Ozone concentrations in GRSM have 

14 been among the highest in the eastern U.S. In the recent conditions scenario, the grid cells 

15 representing 44 percent of GRSM had three year average W126 value above 15 ppm-hrs. After

16 adjustments to just meet the current standard, no grid cell had a three-year average W126 value 

17 above 7 ppm-hrs. In the recent conditions scenario for ROMO, three-year average W126 values 

18 for all grid cells were above15 ppm-hrs. In the current standard scenario, values for 59 percent 
5-51 

1 of the park were below 7 ppm-hrs. For SEKI, three-year average W126 values for all grid cells 

2 were above 15 ppm-hrs in the recent conditions scenario, but dropped below 7 ppm-hrs for the 

3 current standard scenario (U.S. EPA, 2014, section 7.4). 

4 In summary, these analyses indicate that O3 concentrations in U.S. national parks in 

5 recent years correspond to W126 index values at which some foliar injury may occur, with 

6 variation associated with relative soil moisture conditions. None of the 214 parks assessed are 

7 estimated to exceed the annual benchmark criteria for the base scenario (W126 >10.46 ppm-hrs) 

8 after adjusting air quality to meet the current standard. Although model-adjusted scenarios to 

9 just meet the current standard indicate substantial reductions in three-year average W126 index 

10 values estimated by the VNA approach, some individual year values may range higher. The case 

11 study analysis of three parks indicates the potential for appreciable ecosystem services impact 

12 associated with foliar injury, although it might be expected that such impact would relate more to 

13 severe and/or widespread foliar injury occurrences. While these analysis indicate the potential 

14 for foliar injury to occur under conditions that meet the current standard, the extent of foliar 

15 injury that might be expected under such conditions is unclear from these analyses. 

16 • To what extent are the exposures and risks remaining upon simulating just 

17 meeting the current O3 standard important from a public welfare perspective? 

18 The screening level assessment, as described above, indicates that risk of visible foliar 

19 injury is likely to be lower in some national parks after simulating just meeting the current 

20 standard. Based on the national-scale analysis, visible foliar injury would likely continue to 

21 occur at lower O3 exposures, including some sensitive species growing in areas (e.g., National 

22 Parks and other Class I areas) that may provide important cultural ecosystem services to the 

23 public. Staff notes that such occurrences might reasonably be considered to have some 

24 importance from a public welfare perspective, as discussed in section 5.1 above. 

25 • What are the ecosystem services potentially affected by visible foliar injury, to 

26 what degree can the magnitude of these effects be qualitatively or quantitatively 

27 characterized, and to what extent are they important from a public welfare 

28 perspective? 

29 The ecosystem services most likely to be affected by O3-induced foliar injury are cultural 

30 services, including aesthetic value and outdoor recreation. Aesthetic value and outdoor 

31 recreation depend on the perceived scenic beauty of the environment. Many outdoor recreation 

32 activities directly depend on the scenic value of the area, in particular scenic viewing, wildlife33 watching, hiking, and camping. These activities and services are of significant importance to 

34 public welfare as they are enjoyed by millions of Americans every year and generate millions of 

35 dollars in economic value (U.S. EPA, 2014, Chapter 5, Chapter 7). These aesthetic values are at 
5-52 

1 risk of impairment because of O3-induced damage directly due to foliar injury. Other ecosystem 

2 services that have also been found to be associated with O3-sensitive plants include those that fall 

3 under the categories of provisioning. For example, several tribes have indicated that many of the 

4 known confirmed O3-sensitive species (including bioindicator species) are culturally significant 

5 (see Appendix 5A). Although data are not available to explicitly quantify these negative effects

6 on ecosystem services, several qualitative analyses conducted in the WREA are summarized 

below. 7 

8 To assess the effects of visible foliar injury on recreation, the WREA reviewed the 

9 National Survey on Recreation and the Environment (NSRE), as well as the 2006 National 

10 Survey of Fishing, Hunting, and Wildlife-Associated Recreation (FHWAR) and a 2006 analysis 

11 done for the Outdoor Industry Foundation (OIF). According to the NSRE, some of the most 

12 popular outdoor activities are walking, including day hiking and backpacking; camping; bird 

13 watching; wildlife watching; and nature viewing. Participant satisfaction with these activities 

14 can depend on the quality of the natural scenery, which can be adversely affected by O3-related 

15 visible foliar injury. According to the FHWAR and the OIF reports, the total expenditures across 

16 wildlife watching activities, trail-based activities, and camp-based activities are approximately 

17 $230 billion dollars annually. While the WREA could not quantify the magnitude of the impacts 

18 of O3 damage to the scenic beauty and outdoor recreation, the existing losses associated with 

19 current O3-related foliar injury are reflected in reduced outdoor recreation expenditures (U.S. 

20 EPA, 2014, section 7.1). 

21 The WREA also assessed O3 impacts on cultural ecosystem services related to foliar 

22 injury at three national parks – Great Smoky Mountains National Park, Rocky Mountain 

23 National Park, and Sequoia/Kings National Parks - by considering information on visitation 

24 patterns, recreational activities and visitor expenditures. The analysis included percent cover of 

25 species sensitive to foliar injury and focused on the overlap between recreation areas within the 

26 park and elevated W126 concentrations. All three of these park units are in areas that are known 

27 to have high ambient O3 concentrations, have vegetation maps, and have species that are 

28 considered O3 sensitive. Using GIS, the NPS vegetation maps were compared to the national O3

29 surface to illustrate where foliar injury may be occurring, particularly with respect to park 

30 amenities such as trails (U.S. EPA, 2014, section 7.4). 

31 Great Smoky Mountains National Park is prized, in part, for its rich species diversity. 

32 The large mix of species includes 37 O3-sensitive species and many areas contain several 

33 sensitive species. With 3.8 million hikers using the trails every year and those hikers willing to 

34 pay (WTP) over $266 million for that activity, even a small benefit of reducing O3 damage in the 

35 park could result in a significant economic value. Ozone concentrations in Great Smoky 

36 Mountains National Park have been among the highest in the eastern U.S. – at times twice as 
5-53 

1 high as neighboring cities such as Atlanta (U.S. EPA, 2014, p. 7-52). Unlike Great Smoky 

2 Mountains National Park, sensitive species cover in Rocky Mountain National Park is driven by 

3 a few O3-sensitive species (7 species) and most notably by Quaking Aspen. This is significant in 

4 that many of the visitors to Rocky Mountain National Park visit specifically to see this tree in its 

5 fall foliage. Given 1.5 million hikers in Rocky Mountain National Park and their $70 million 

6 WTP for the hiking experience, even a small improvement in the scenic value could be 

7 economically significant (U.S. EPA, 2014, section 7.4.2, p. 7-56). Sequoia/Kings National Parks 

8 is home to 12 identified sensitive species. Again, although the EPA is not able to quantify the 

9 impact of this scenic damage on hiker satisfaction for hikers in Sequoia/Kings National Parks 

10 and their $26 million WTP for the experience, even a small improvement in the scenic value 

11 could be economically significant ((U.S. EPA, 2014, section 7.4.3, p. 7-63).

12 • What are the uncertainties associated with this information and what is the level 

13 of confidence associated with those estimates? 

14 Uncertainties associated with these analyses are discussed in the WREA, sections 7.3.2 

15 and 8.5.3, and in WREA Table 7.23. As discussed in the WREA (section 8.5.3), evaluating soil 

16 moisture is more subjective than evaluating O3 exposure because of its high spatial and temporal 

17 variability within the O3 season, and there is considerable subjectivity in the categorization of 

18 relative drought. The WREA generally concludes that the spatial and temporal resolution for the 

19 soil moisture data is likely to underestimate the potential of foliar injury that could occur in some 

20 areas. In addition, there is lack of a clear threshold for drought below which visible foliar injury 

21 would not occur. In general, low soil moisture reduces the potential for foliar injury, but injury 

22 could still occur, and the degree of drought necessary to reduce potential injury is not clear. Due 

23 to the absence of biosite injury data in the Southwest region and limited biosite data in the West 

24 and West North Central regions, the benchmarks applied may not be applicable to these regions. 

25 There are also important uncertainties in the estimated O3 concentrations for the different 

26 air quality scenarios evaluated (U.S. EPA, 2014, section 8.5), as discussed earlier in this section. 

27 In general, this interpolation method under-predicts higher 12-hr W126 exposures. Due to the 

28 important influence of higher exposures in determining risks to plants, the potential for the VNA 

29 interpolation approach to under-predict higher W126 exposures could result in an under30 estimation of risks to vegetation in some areas. The WREA applied the benchmarks from the 

31 national-scale analysis to a screening-level analysis of 214 national parks and case studies of 

32 three national parks. Therefore, uncertainties in the foliar injury benchmarks and in the air 

33 quality analyses are propagated into the national park analyses. The uncertainties associated 

34 with air quality assessments include those resulting from use of an unevenly distributed 

35 monitoring network with fewer monitors in rural and western sites to drive a VNA interpolation 
5-54 

1 and the use of model-adjusted air quality to generate alternative air quality scenarios including 

2 just meeting the current standard (these uncertainties are described in more detail above). 

3 Additional uncertainties are associated with the national park analyses. Specifically, 

4 survey estimates of participation rates, visitor spending/economic impacts, and willingness-to5 pay are inherently uncertain. These surveys potentially double-count impacts based on the 

6 allocation of expenditures across activities but also potentially exclude other activities with 

7 economic value. Each survey uses different survey methods, so it is not appropriate to generalize 

8 across the surveys. In general, the national level surveys apply standard approaches, which 

9 minimize potential bias. Other sources of uncertainty are associated with the mapping, including 

10 park boundaries, vegetation species cover, and park amenities, such as scenic overlooks and 

11 trails. In general, the WREA concludes that there is high confidence in the park mapping (U.S. 

12 EPA, 2014, Table 7-23).. 

13 5.5 OTHER WELFARE EFFECTS 

14 In addition to the welfare effects discussed in the previous sections, there is evidence of 

15 other O3 effects, such as those related to climate impacts that we consider here. In this section, 

16 the WREA national-scale analyses of the effects of insect damage to forests related to elevated 

17 O3 exposures are considered in section 5.5.1, and a case study-scale characterization of the effect 

18 community composition changes on forest susceptibility and fire regulation in California is 

19 considered in section 5.5.2. As above, these sections, where possible, consider the WREA 

20 information regarding risk remaining under model adjusted conditions just meeting the current 

21 standard and associated uncertainties (U.S. EPA 2014, section 8.5). Chapters 5, 6, and 7 of the 

22 WREA also qualitatively assessed additional ecosystem services, including regulating services 

23 such as hydrologic cycle and pollination; provisioning services such as commercial non-timber 

24 forest products; and cultural services with aesthetic and non-use values. The information 

25 associated with these latter effects is insufficient to inform the target protection of the standard. 

26 The effects of O3 on climate are also considered in section 5.5.3 below, drawing primarily on the 

27 evidence presented in the ISA (U.S. EPA 2013, chapter 10). 

28 5.5.1 Forest Susceptibility to Insect Infestation

29 Ozone in ambient air can contribute to increased susceptibility of some forests to 

30 infestation by some chewing insects, including the southern pine beetle and western bark beetle 

31 (U.S. EPA 2013, chapter 9; U.S. EPA 2014, sections 5.3.3 and 5.4). These infestations can cause 

32 economically significant damage to tree stands and the associated timber production. The 

33 WREA described the potential impacts of this effect on timber markets (U.S. EPA 2014, section 

34 5.4). In the short-term, the immediate increase in timber supply that results from the additional
5-55 

1 harvesting of damaged timber depresses prices for timber and benefits consumers. In the longer2 term, the decrease in timber available for harvest raises timber prices, potentially benefitting

3 producers. The United States Forest Service (USFS) reports timber producers have incurred 

4 losses of about $1.4 billion (2010$), and wood-using firms have gained about $966 million, due 

5 to beetle outbreaks between 1977 to 2004. It is not possible to attribute a portion of these 

6 impacts resulting from the effect of O3 on trees’ susceptibility to insect attack; however, the 

7 losses are embedded in the estimates cited. 

8 To provide some quantitative estimates related to insect infestation-related risks, the 

9 WREA reported the estimates of 3-year average W126 values in areas estimated to be at risk of 

10 greater than 25% timber loss (high loss) due to pine beetle infestation. This was done for all six 

11 WREA air quality scenarios. For example, for the recent conditions scenario, approximately 57 

12 percent of the at-risk area has W126 estimates above 15 ppm-hrs, with the percentage dropping 

13 to approximately five percent in the current standard scenario (U.S. EPA 2014, section 5.4). 

14 5.5.2 Fire Regulation 

15 Evidence indicates that fire regime regulation may also be negatively affected by O3

16 exposure (U.S. EPA 2013, chapter 9; U.S. EPA 2014, section 5.3.3). For example, Grulke et al. 

17 (2008) reported various lines of evidence indicating that O3 exposure may contribute to southern 

18 California forest susceptibility to wildfires by increasing leaf turnover rates and litter, increasing 

19 fuel loads on the forest floor. According to the National Interagency Fire Center, in the U.S. in 

20 2010 over 3 million acres burned in wildland fires and an additional 2 million acres were burned 

21 in prescribed fires. From 2004 to 2008, Southern California alone experienced, on average, over 

22 4,000 fires per year burning, on average, over 400,000 acres per fire. The California Department 

23 of Forestry and Fire Protection (CAL FIRE) estimated that losses to homes due to wildfire were 

24 over $250 million in 2007 (CAL FIRE, 2008). In 2008, CAL FIRE’s costs for fire suppression 

25 activities were nearly $300 million (CAL FIRE, 2008). 

26 The WREA developed maps that overlay the mixed conifer forest area of California with 

27 areas of moderate or high fire risk defined by CAL FIRE and with recent W126 concentrations 

28 and surfaces adjusted to just meet existing and alternative standards. The highest fire risk and 

29 highest O3 concentrations overlap with each other, as well as with significant portions of mixed 

30 conifer forest. In the recent concentrations scenario, over 97 percent of mixed conifer forest area 

31 has average W126 values over 7 ppm-hrs with a moderate to severe fire risk, and 74 percent has 

32 average W126 values over 15 ppm-hrs with a moderate to severe fire risk. The scenario for air 

33 quality adjusted to just meet the current standard, almost all of the mixed conifer forest area with 

34 a moderate to high fire risk shows a reduction in O3 to below a W126 index value of 7 ppm-hrs

35 (average across three years of scenario). In the scenario for an average W126 index value of 15 
5-56 

1 ppm-hrs, all but 0.18 percent of the area has average index values below 7 ppm-hrs, and for the 

2 W126 scenarios of 11 and 7 ppm-hrs, all of the moderate to high fire threat area has estimated 

3 average W126 values below 7 ppm-hrs (U.S. EPA 2014, section 5.3.3, figure 5-3). 

4 5.5.3 Ozone Effects on Climate

5 Tropospheric O3 is a major greenhouse gas, third in importance after carbon dioxide 

6 (CO2) and methane (CH4). While the developed world has successfully reduced emissions of O3

7 precursors in recent decades, many developing countries have experienced large increases in 

8 precursor emissions and these trends are expected to continue, at least in the near term (U.S. 

9 EPA 2013, section 10.3.6.2). Projections of radiative forcing due to changing O3 over the 21st 

10 century show wide variation, due in large part to the uncertainty of future emissions of source 

11 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 12 , depending on the emissions scenario (U.S. 

13 EPA 2013, section 10.3.6.2; Stevenson et al., 2006). Reduction of tropospheric O3

14 concentrations could therefore provide an important means to slow climate change in addition to 

15 the added benefit of improving surface air quality (U.S. EPA, 2013, section 10.5). 

16 It is clear that increases in tropospheric O3 lead to warming. However the precursors of 

17 O3 also have competing effects on the greenhouse gas CH4, complicating emissions reduction 

18 strategies. A decrease in CO or VOC emissions would enhance OH concentrations, shortening 

19 the lifetime of CH4, while a decrease in NOX emissions could depress OH concentrations in 

20 certain regions and lengthen the CH4 lifetime (U.S. EPA, 2013, section 10.5). 

21 Abatement of CH4 emissions would likely provide the most straightforward means to 

22 address O3-related climate change since CH4 is itself an important precursor of background O3

23 (West et al., 2007; West et al., 2006; Fiore et al., 2002). A reduction of CH4 emissions would 

24 also improve air quality in its own right. A set of global abatement measures identified by West 

25 and Fiore (2005) could reduce CH4 emissions by 10% at a cost savings, decrease background O3

26 by about 1 ppb in the Northern Hemisphere summer, and lead to a global net cooling of 0.12 

W/m2 27 . West et al. (2007) explored further the benefits of CH4 abatement, finding that a 20% 

28 reduction in global CH4 emissions would lead to greater cooling per unit reduction in surface O3, 

29 compared to 20% reductions in VOCs or CO (U.S. EPA, 2013, section 10.5). 

30 Important uncertainties remain regarding the effect of tropospheric O3 on future climate 

31 change. To address these uncertainties, further research is needed to: (1) improve knowledge of 

32 the natural atmosphere; (2) interpret observed trends of O3 in the free troposphere and remote 

33 regions; (3) improve understanding of the CH4 budget, especially emissions from wetlands and 

34 agricultural sources, (4) understand the relationship between regional O3 radiative forcing and 
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1 regional climate change; and (5) determine the optimal mix of emissions reductions that would 

2 act to limit future climate change (U.S. EPA, 2013, section 10.5). 

3 The IPCC has estimated the effect of the tropospheric O3 change since preindustrial times 

4 on climate has been estimated to be about 25-40% of the anthropogenic CO2 effect and about 

5 75% of the anthropogenic CH4 effect. There are large uncertainties in the radiative forcing 

6 estimate attributed to tropospheric O3, making the effect of tropospheric O3 on climate more 

7 uncertain than the effect of the long-lived greenhouse gases (U.S. EPA, 2013, section 10.5). 

8 Radiative forcing does not take into account the climate feedbacks that could amplify or 

9 dampen the actual surface temperature response. Quantifying the change in surface temperature 

10 requires a complex climate simulation in which all important feedbacks and interactions are 

11 accounted for. As these processes are not well understood or easily modeled, the surface 

12 temperature response to a given radiative forcing is highly uncertain and can vary greatly among 

13 models and from region to region within the same model (U.S. EPA, 2013, section 10.5). 

14 5.5.4 Additional Effects

15 Recent information available since the last review considers the effects of O3 on chemical 

16 signaling in insect and wildlife interactions. Specifically, studies on O3 effects on pollination and 

17 seed dispersal, defenses against herbivory and predator-prey interactions all consider the ability 

18 of O3 to alter the chemical signature of VOCs emitted during these pheromone-mediated events. 

19 The effects of O3 on chemical signaling between plants, herbivores and pollinators as well as 

20 interactions between multiple trophic levels is an emerging area of study that may result in 

21 further elucidation of O3 effects at the species, community and ecosystem-level (U.S. EPA, 2013, 

22 p. 9-98). 

23 5.6 CASAC ADVICE

24 Following the 2008 decision to revise the secondary standard by setting it identical to the 

25 revised primary standard, CASAC conveyed additional advice to the Administrator regarding 

26 that decision. Shortly after that, several petitioners filed suit challenging the decision and in 

27 September 2009, the EPA announced its intention to reconsider the 2008 standards, issuing a 

28 notice of proposed rulemaking in January 2010 (FR 75 2938). Soon after, the EPA solicited 

29 CASAC review of that proposed rule and in January 2011 solicited additional advice. This 

30 proposal was based on the scientific and technical record from the 2008 rulemaking, including 

31 public comments and CASAC advice and recommendations. As further described in section 

32 1.2.2 above, the EPA in the fall of 2011 did not promulgate final rulemaking in that process but 

33 decided to coordinate further proceedings on the reconsideration rulemaking with this ongoing 
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1 periodic review. Accordingly, we are describing CASAC advice related to the 2008 final 

2 decision and the subsequent reconsideration in this section, as well as advice on the NAAQS 

3 review that was initiated in September 2008. 

4 In April 2008, the members of the CASAC Ozone Review Panel sent a letter to EPA 

5 stating that “[i]n our most-recent letters to you on this subject - dated October 2006 and March 

6 2007 - ... the Committee recommended an alternative secondary standard of cumulative form 

7 that is substantially different from the primary Ozone NAAQS in averaging time, level and form 

8 — specifically, the W126 index within the range of 7 to 15 ppm-hours, accumulated over at least 

9 the 12 ‘daylight’ hours and the three maximum ozone months of the summer growing season ” 

10 (Henderson, 2008). The letter continued: 

11 The CASAC now wishes to convey, by means of this letter, its additional, 

12 unsolicited advice with regard to the primary and secondary Ozone NAAQS. In 

13 doing so, the participating members of the CASAC Ozone Review Panel are 

14 unanimous in strongly urging you or your successor as EPA Administrator to 

15 ensure that these recommendations be considered during the next review cycle for 

16 the Ozone NAAQS that will begin next year ... The CASAC was also greatly 

17 disappointed that you failed to change the form of the secondary standard to 

18 make it different from the primary standard. As stated in the preamble to the 

19 Final Rule, even in the previous 1996 ozone review, “there was general 

20 agreement between the EPA staff, CASAC, and the Administrator, ... that a 

21 cumulative, seasonal form was more biologically relevant than the previous 1-

22 hour and new 8-hour average forms (61 FR 65716)” for the secondary 

23 standard......Unfortunately, this scientifically-sound approach of using a 

24 cumulative exposure index for welfare effects was not adopted... 

25 In response to the EPA’s solicitation of their advice on the Agency’s proposed 

26 rulemaking as part of the reconsideration, CASAC conveyed their support as follows (Samet, 

27 2010). 

28 CASAC also supports EPA’s secondary ozone standard as proposed: a new 

29 cumulative, seasonal standard expressed as an annual index of the sum of 

30 weighted hourly concentrations (i.e., the W126 form), cumulated over 12 hours 

31 per day (8am to 8pm) during the consecutive 3-month period within the ozone 

32 season with the maximum index value, set as a level within the range of 7 to [1]5 

33 ppm-hours. This W126 metric can be supported as an appropriate option for 

34 relating ozone exposure to vegetation responses, such as visible foliar injury and 

35 reductions in plant growth. We found the Agency’s reasoning, as stated in the 

36 Federal Register notice of January 19, 2010, to be supported by the extensive 

37 scientific evidence considered in the last review cycle. In choosing the W126 form 

38 for the secondary standard, the Agency acknowledges the distinction between the 

39 effects of acute exposures to ozone on human health and the effects of chronic 

40 ozone exposures on welfare, namely that vegetation effects are more dependent on 
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1 the cumulative exposure to, and uptake of, ozone over the course of the entire 

2 growing season (defined to be a minimum of at least three months). 

3 In advice offered so far in the current review, which is considering an updated scientific 

4 and technical record since the 2008 rulemaking, CASAC indicated that a conclusion that the 

5 current standard is inadequate to protect vegetation and ecosystems is “warranted” although the 

6 foundation needs to be broadened beyond analysis focused on Class I areas and trees to include 

7 “effects on sensitive crops, trees in regions outside of Class I areas, and additional ecosystem 

8 impacts” (Frey and Samet, 2012b, p. 2). The Panel additionally endorsed the first draft PA 

9 discussions and conclusions on biologically relevant exposure metrics, stating that “the focus on 

10 the W126 form is appropriate” and that “there is a strong justification made for using a 

11 cumulative and weighted exposure standard for welfare effects (i.e. the W126)...” (Frey and 

12 Samet, 2012b, p. 2). 

13 5.7 PRELIMINARY STAFF CONCLUSIONS ON ADEQUACY OF 

14 SECONDARY STANDARD

15 This section presents preliminary staff conclusions for the Administrator to consider in 

16 deciding whether the existing secondary O3 standard is adequate and whether it should be 

17 retained or revised. Our conclusions are based on consideration of the assessment and integrative 

18 synthesis of information presented in the ISA, as well as our analyses of air quality distributions; 

19 analyses in the second draft WREA; and the comments and advice of CASAC and public 

20 comment on an earlier draft of this document and the ISA and WREA, as discussed above. 

21 Taking into consideration the responses to specific questions discussed above, we revisit the 

22 overarching policy question for this chapter: 

23 • Does the currently available scientific evidence and exposure/risk information, as 

24 reflected in the ISA and REA, support or call into question the adequacy and/or 

25 appropriateness of the protection afforded by the current secondary O3 standard?

There is a longstanding evidence base on the phytotoxic effects of O3. 26 Since first 

27 identified by Richards, et al in 1958, numerous studies have been conducted on the effects of O3

28 on plants and their associated ecosystems. Taken together, these studies demonstrate that O3 -

29 induced effects that occur at the subcellular and cellular levels, at sufficient magnitudes 

30 propagate up to produce larger scale effects that affect the whole organism. In particular in this 

31 review, many of the recent studies have focused on and further increased our understanding of 

32 the molecular, biochemical and physiological mechanisms that explain how plants are affected 

33 by O3, in the absence of other stressors, particularly in the area of genomics (U.S. EPA, 2013, 

34 Chapter 9, section 9.3). 
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1 Based on its assessment of this extensive body of science, the ISA determined that a 

2 causal relationship exists between exposure to O3 in ambient air and visible foliar injury effects 

3 on vegetation, reduced vegetation growth, reduced productivity in terrestrial ecosystems, reduced 

4 yield and quality of agricultural crops and alteration of below-ground biogeochemical cycles 

5 (U.S. EPA 2013, Table 1-2). Additionally, the ISA determined that a likely to be causal 

6 relationship exists between exposures to O3 in ambient air and reduced carbon sequestration in 

7 terrestrial ecosystems, alteration of terrestrial ecosystem water cycling and alteration of 

8 terrestrial community composition (U.S. EPA, 2013, Table 1-2). 

9 Recent studies also continue to provide evidence that adverse vegetation effects are

10 attributable to cumulative seasonal O3 exposures. On the basis of the entire body of evidence in 

11 this regard, the ISA concludes that “quantifying exposure with indices that cumulate hourly O3

12 concentrations and preferentially weight the higher concentrations improves the explanatory 

13 power of exposure/response models for growth and yield, over using indices based on mean and 

14 peak exposure values” (U.S. EPA, 2013, p. 2-44). Thus, as in other recent reviews, the evidence

15 continues to provide a strong basis for concluding that it is appropriate to judge impacts of O3 on 

16 vegetation, related effects and services, and the level of public welfare protection achieved, using 

17 a cumulative, seasonal exposure metric, such as the W126-based metric. In addition, CASAC 

18 has consistently since the 1997 review expressed support for the use of such a metric, including 

19 most recently in its letter on the first draft PA (Frey and Samet, 2012b, p. 2). Thus, based on the 

20 long-established conclusions and long-standing supporting evidence described above, we 

21 continue to conclude that the most appropriate and biologically-relevant way to relate O3

22 exposure to plant growth, and to determine what would be adequate protection for public welfare 

23 effects attributable to the presence of O3 in the ambient air, is to express exposures in terms of a 

24 cumulative seasonal form, and in particular the W126 metric. Accordingly, in considering the 

25 current evidence and exposure/risk information with regard to the adequacy of public welfare 

26 protection it affords, staff has considered both the evidence of vegetation and welfare impacts in 

27 areas of the U.S. likely to have met the current standard, as well as air quality information 

28 regarding W126 index values (and evidence of effects associated with those values) in areas 

29 likely to have met the current standard. 

30 We further take note of CASAC advice in the last review, where they have recognized 

31 that the current secondary standard is of an averaging time and form pertaining more to O3

32 effects on human health than to the effects of O3 exposure on welfare. In so doing, they 

33 emphasized the dependency of vegetation effects on the cumulative exposure to and uptake of O3

34 over the course of the entire growing season (Henderson, 2006). Accordingly, we recognize 
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1 potential uncertainty in the level of protection provided by the current standard from vegetation 

2 related effects, due to its averaging time and form. 

3 Based on the considerations described in the sections above and summarized below, we

4 preliminarily conclude that the currently available evidence and exposure/risk information calls 

5 into question the adequacy of public welfare protection provided by the current standard and 

6 provides support for considering potential alternative standards to achieve greater public welfare 

7 protection, especially for sensitive vegetation and ecosystems in federally protected Class I and 

8 similar areas. 

9 As an initial matter, we note that the CAA does not require that a secondary standard be 

10 protective of all effects associated with a pollutant in the ambient air, but only those considered 

11 adverse to the public welfare (as described in section 1.3.2 above). In judging whether particular 

12 known or anticipated effects should be considered adverse, the Administrator considers a number 

13 of factors, including the intended use of the affected entity(s) and the location(s) in which those 

14 effects occur or are predicted to occur. In this regard, effects that have been observed in 

15 federally protected Class I areas which are afforded special protections under the CAA may have 

16 greater significance to the public welfare. In this context, staff has given particular consideration 

17 to recent study findings of welfare-related effects in such areas, in combination with information 

18 on air quality during the study periods to inform our conclusions regarding the nature and scope 

19 of welfare effects in such protected areas, including areas where the current standard is met or 

20 likely to have been met. 

21 Regarding visible foliar injury, new research includes: 1) controlled exposure studies; 2) 

22 multi-year field surveys; 3) US FS FIA biomonitoring program surveys and assessments. In 

23 addition to supporting the ISA causal determination, these studies also address some 

24 uncertainties indentified in the previous review (i.e., influence of soil moisture on visible injury 

25 development) and further show that visible foliar injury can occur under conditions where 8-hour 

26 average O3 concentrations are or would be expected to be below the level of the current standard 

27 (e.g., Kline et al., 2008, as discussed in section 5.4.1 above). Incidence of visible foliar injury 

28 symptoms on O3-sensitive vegetation has also been documented in the field in federally 

29 protected areas that have met the current standard. As one example, in Cape Romain National 

30 Wildlife Refuge in SC, where air quality data indicate O3 concentrations have met the current 

31 standard over the period from 2001-2012, recent study findings indicate that three species have 

32 exhibited O3-induced visible foliar injury (winged sumac, Chinese tallow tree, and wild grape). 

33 In at least one of the studied locations, up to 32 % of the wild grape plants examined showed 

34 foliar injury symptoms during years where 8-hour O3 concentrations were similar to the level of 

35 the current standard (U.S. EPA 2013, section 9.4.2.1). Importantly, these O3-induced vegetation 
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1 effects were found in Class I areas which have identified O3-impacts to vegetation as an air 

2 quality related value that should be protected commiserate with its congressionally recognized 

3 public welfare significance (75 FR 3024). A staff air quality assessment identified Class I areas

4 with recent air quality that met the current standard but was above a W126 index value of 15 

5 ppm-hrs (Table 5-2 above). Given evidence of the potential occurrence of visible foliar injury at 

6 W126 index values of this magnitude, staff concludes that air quality levels that are at or below 

7 the level of the current standard may allow levels of visible foliar injury incidence to occur in 

8 areas of special significance to the public welfare. 

9 Results from the exposure and risk assessments also indicate visible foliar injury at sites 

10 with air quality likely to meet the current standard. In the national-scale analysis of the presence 

11 or absence of foliar injury across the years 2006-2010 in U.S. National Forest sites with annual 

12 W126 index values of from 7 to 15 ppm-hrs, the percent of sites indicating the presence of foliar 

13 injury generally ranges from about 5 to 20 percent, varying with soil moisture conditions for the 

14 year. Across all sites, regardless of the year’s soil moisture conditions, these analyses find an 

15 increase in the proportion of sites showing the presence of foliar injury or elevated foliar injury 

16 with increasing W126 index values up to W126 values of approximately 10 ppm-hrs (see U.S. 

17 EPA 2014, Figure 7-10). Using this relationship, a National Park screening assessment (214 

18 parks) found that 58% of the parks analyzed exceeded a W126 of 10.46 ppm-hrs, which is 

19 associated with 17.7% injury prevalence. Data indicate large reductions in injury prevalence 

20 with incrementally lower W126 index values. Just meeting the current standard would bring all 

21 parks under 10.46 ppm-hrs. However, as noted above, there are important uncertainties 

22 associated with the model-adjusted air quality used in these assessments (U.S. EPA, 2014, Table 

23 7-23, section 8.5.2) that we note in considering what the adjustment results convey with regard to 

24 W126 conditions associated with just meeting the current standard. 

25 We also note that visible foliar injury is associated with important cultural and 

26 recreational ecosystem services to the public, such as scenic viewing, wildlife-watching, hiking, 

27 and camping, that are of significance to the public welfare and enjoyed by millions of Americans 

28 every year, generating millions of dollars in economic value (U.S. EPA 2014, section 7.1). In 

29 addition, several tribes have indicated that many of the known confirmed O3-sensitive species 

30 (including bioindicator species) are culturally significant (see Appendix 5A). These ecosystem

31 services are at risk of impairment because of O3-induced damage directly due to foliar injury, 

32 though data is not available to explicitly quantify these negative effects. Thus, while it is unclear 

33 to what extent O3-induced visible foliar injury impacts these services, we conclude it is 

34 appropriate to consider the potential for these effects to contribute to adverse effects on public 
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1 welfare in considering the adequacy and appropriateness of the protection afforded by the current 

2 standard. 

3 Tree biomass loss is another O3-induced effect that is important from a public welfare 

4 perspective, both alone and in conjunction with associated ecosystem-level impacts. This effect 

5 occurs in both Class I and non-Class I areas and its importance comes from whether this effect 

6 reduces the intended use of the plant or its associated ecosystems in a way that is relevant to the 

7 public welfare. Recent studies confirm and extend the evidence of O3-related effects on tree 

8 growth, productivity and carbon storage. Analysis of existing data conducted by the EPA staff 

9 and discussed in the ISA has substantially reduced the uncertainty associated with using OTC C10 R functions to predict tree growth effects in the field, as described in section 5.2.1 above (U.S. 

11 EPA, 2013, section 9.6.3.2). Based on the 12 individual tree seedling C-R functions, seven of 

12 the 12 species show 2% seedling biomass loss at a W126 index value below 8 ppm-hrs and in the 

13 other five species at a W126 value above 18 ppm-hrs. Within the group of seven more sensitive 

14 species, the most sensitive are cottonwood and black cherry (U.S. EPA 2014, section 6.2). 

15 Based on the median composite C-R functions, the W126 index values associated with a 2% 

16 biomass loss range from approximately 7 to 14 ppm-hrs (U.S. EPA, 2014, Figure 6-4, section 

17 6.2.1.2). Recent studies have provided additional evidence on tree biomass or growth effects 

18 associated with multiple year exposures in the field, including the potential for compounding and 

19 carry-over effects. These effects could lead to a negative impact on species regeneration and 

20 have implications for the subsequent growing season in the following year (U.S. EPA, 2013, 

21 section 9.4), as recognized in recent advice from CASAC (Frey and Samet, 2012b). 

22 Accordingly, in considering WREA biomass loss estimates, we also consider the WREA 

23 evaluation of a potential for underestimation of such effects (U.S. EPA, 2014, section 6.2.1.4). 

24 While it is not possible at this time to identify the extent or magnitude of such effects in the field 

25 under exposure levels that may be associated with the current standard, their occurrence, when of 

26 a magnitude of concern, on federally protected lands or any associated impacts on ecosystem 

27 services such as habitat quality, hydrologic regimes, carbon storage or air pollution removal in 

28 areas important to the public might reasonably be concluded to be significant to public welfare. 

29 We additionally take note of the WREA findings with regard to urban areas where tree CO2

30 sequestration and air pollutant removal are important public welfare services. 

31 With respect to crops, the detrimental effect of O3 on crop production has been 

32 recognized since the 1960s, and recent O3 concentrations in many areas across the U.S. are high 

33 enough that they might be expected to cause yield loss in a variety of agricultural crops 

34 including, but not limited to, soybeans, wheat, potatoes, watermelons, beans, turnips, onions, 

35 lettuces, and tomatoes (U.S. EPA, 2013, section 9.4.4). In general, the vast majority of the new 

36 scientific information confirms prior conclusions that exposure to O3 can decrease growth and 
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1 yield of crops. Recent research has highlighted the effects of O3 on crop quality. Increasing O3

2 concentration decreases nutritive quality of grasses, decreases macro- and micro-nutrient 

3 concentrations in fruits and vegetable crops (U.S. EPA 2013, section 9.4.4). Recent studies 

4 continue to find yield loss levels in crop species studied previously under NCLAN that reflect 

5 the earlier findings. There has been little published evidence that crops are becoming more 

6 tolerant of O3 (U.S. EPA, 2006a; U.S. EPA 2013). This is especially evident in the research on 

7 soybean. The 2013 ISA reported comparisons between yield predictions based on data from 

8 cultivars used in NCLAN studies, and yield data for modern cultivars from SoyFACE (U.S. 

9 EPA, 2013, section 9.6.3). They confirm that the average response of soybean yield to O3

10 exposure has not changed in current cultivars. In addition, satellite and ground-based O3 

11 measurements have been used to assess yield loss caused by O3 over the continuous tri-state area 

12 of Illinois, Iowa, and Wisconsin. The results showed that O3 concentrations reduced soybean 

13 yield, which correlates well with the previous results from FACE- and OTC-type experiments 

14 (U.S. EPA, 2013, section 9.4.4.1). Thus, the recently available evidence, as assessed in the ISA, 

15 continues to support the conclusions of the 1996 and 2006 Criteria Documents that ambient O3

16 concentrations can reduce the yield of major commodity crops in the U.S. 

17 The currently available evidence, as assessed in the ISA, continues to support the use of 

18 C-R functions based on OTC experiments. Further, important uncertainties have been reduced 

19 regarding the exposure-response functions for crop yield loss, especially for soybean. In general, 

20 the ISA reports consistent results across exposure techniques and across crop varieties (U.S. EPA 

21 2013, section 9.6.3.2). Based on the crop C-R functions, the W126 index values associated with 

22 a five percent yield loss range from approximately 12 to 17 ppm-hrs. The lower end of the range 

23 would be more protective of soybeans which are the second-most planted field crop in the U.S. 

24 (http://www.ers.usda.gov/topics/crops/soybeans-oil-crops/background.aspx). Staff analyses of 

25 recent monitoring data (2009-2011) indicate that O3 concentrations in multiple agricultural areas 

26 in the U.S. that meet the current standard correspond to W126 index levels above 12 ppm-hrs. 

27 The information for other welfare effects, including those with causal or likely causal 

28 relationships with O3 (i.e., alteration of ecosystem water cycling, changes in climate), is limited 

29 with regard to our ability to consider potential impacts of air quality conditions associated with 

30 the current standard, although the WREA provides some perspective on this issue with regard to 

31 susceptibility to insect attack and fire regime change. We note, however, the importance of these 

32 effects categories to the public welfare.

33 Given all of the above, staff reaches the preliminary conclusion that the available 

34 evidence and exposure and risk information calls into question the adequacy of public welfare

35 protection provided by the current standard, and provides strong support to giving consideration 
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1 to revising the current secondary standard to provide increased public welfare protection. More 

2 specifically, staff preliminarily concludes that it is appropriate for the Administrator to consider 

3 revision of the current secondary O3 standard to increase protection against O3-attributable tree 

4 biomass loss, crop yield loss, and visible foliar injury, and particularly for those effects 

5 associated with cumulative, seasonal exposures that occur in specially protected natural areas. In 

6 reaching conclusions on options for the Administrator’s consideration, we note that the final 

7 decision to retain or revise the current secondary O3 standard is a public welfare policy judgment 

8 to be made by the Administrator as to what standards would be requisite (i.e., neither more nor 

9 less stringent than necessary) to protect the public welfare from any known or anticipated 

10 adverse welfare effects. This final decision will draw upon the available scientific evidence for 

11 O3-attributable welfare effects and on analyses of vegetation and ecosystem exposures and public 

12 welfare risks based on impacts to vegetation, ecosystems and their associated services, including 

13 judgments about the appropriate weight to place on the range of uncertainties inherent in the 

14 evidence and analyses. In determining the requisite level of protection for crops and trees, the 

15 Administrator will need to weigh the importance of the predicted risks of these effects in the 

16 overall context of public welfare protection, along with a determination as to the appropriate 

17 weight to place on the associated uncertainties and limitations of this information. 

18 As noted in section 1.3.2 above, our general approach to informing these judgments, 

19 recognizes that the available welfare effects evidence reflects a continuum consisting of higher 

20 O3 concentrations at which scientists generally agree that welfare effects are likely to occur,

21 through lower concentrations at which the likelihood and magnitude of a response become 

22 increasingly uncertain. Therefore, in presenting conclusions here, we are mindful that the 

23 Administrator’s ultimate judgment on the adequacy of protection afforded by the current 

24 standard will most appropriately reflect an interpretation of the available scientific evidence and 

25 exposure/risk information that neither overstates nor understates the strengths and limitations of 

26 that evidence and information. 

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10 nt.org/documents/document.766.aspx.pdf.

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17 Oxidant stipple on grape. Agron. J. 50:559-561.

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19 SRS-118. Asheville, N.C.: US Department of Agriculture, Forest Service, Southern Research Station, pp. 

20 25.

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22 surveys of ozone sensitive plants in northeastern forests (1994-2000). Environ Monit Assess 87: 271-291.

23 Smith, G. (2012). Ambient ozone injury to forest plants in Northeast and North Central USA: 16 years of 

24 biomonitoring. Environ Monit Assess 184: 4049-4065. http://dx.doi.org/10.1007/s10661-011-2243-z 

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29 Sanderson, MG; Savage, NH; Shindell, DT; Strahan, SE; Sudo, K; Szopa, S. (2006). Multimodel ensemble 

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33 human well-being: A framework for assessment. Washington, DC: Island Press.

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40 Standards, Health and Environmental Impacts Division,Research Triangle Park, NC 27711. EPA-452/R41 12-005. Available at: http://www.epa.gov/ttn/naaqs/standards/pm/s_pm_2007_ria.html
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4 Research Triangle Park, NC.

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6 Environmental Protection Agency, Washington, DC. EPA/600/R-05/004aF-cF. March 2006. Available at: 

7 http://www.epa.gov/ttn/naaqs/standards/ozone/s_o3_cr_cd.html

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17 78/004). Washington, DC.

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19 with methane emission controls. PNAS 103: 3988-3993. http://dx.doi.org/10.1073/pnas.0600201103

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21 radiative forcing consequences of changes in ozone precursor emissions. Geophys Res Lett 34: L06806. 

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26 3040.2007.01717.x 

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28 future tropospheric ozone on tree biomass, growth, physiology and biochemistry: A quantitative meta29 analysis. Global Change Biol 15: 396-424. http://dx.doi.org/10.1111/j.1365-2486.2008.01774.x

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31 report. Washington, DC: U.S. Environmental Protection Agency, Science Advisory Board; report no. 

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1 6 CONSIDERATION OF ALTERNATIVE SECONDARY 

2 STANDARDS

3 To the extent that the information available in this review suggests that revision of the 

4 current secondary standard is appropriate to consider, as discussed in chapter 5, staff has 

5 evaluated the available body of evidence, and exposure, risk and air quality information with 

6 regard to the support for consideration of alternative standards, as articulated by the following 

7 overarching question: 

8 • What alternative secondary standards are supported by the currently available 

9 scientific evidence, exposure/risk information and air quality analyses? 

10 To assist us in interpreting the currently available scientific evidence and the results of 

11 recent quantitative exposure/risk analyses to address this question, we have focused on a series 

12 of more specific questions in sections 6.1, 6.2 and 6.3 below. We consider both the scientific 

13 and technical information available at the time of the last review and information newly available 

14 since the last review which has been critically analyzed and characterized in the ISA. 

15 Specifically, we consider the currently available scientific evidence and technical information in 

16 the context of decisions regarding the basic elements of the NAAQS: indicator, (section 6.1); 

17 averaging time and form (section 6.2); and level (section 6.3). CASAC advice on potential 

18 alternative standards is described in section 6.4 and preliminary staff conclusions on potential 

19 alternative standards are discussed in section 6.5. Section 6.6 summarizes preliminary staff 

20 conclusions on adequacy of the current standard and potential alternative standards appropriate 

21 to consider. Key uncertainties in this review and areas in which future research and data 

22 collection would better inform the next review are identified in section 6.7. 

23 6.1 INDICATOR

24 With regard to the indicator for potential alternative secondary standards, we consider the 

25 following question. 

26 • Does the information available in this review continue to support O3 as the

27 indicator for ambient air photochemical oxidants? 

28 In the last review of the air quality for O3 and other photochemical oxidants and of the O3

29 standard, as in other prior reviews, the EPA focused on a standard for O3 as the most appropriate 

30 surrogate for ambient photochemical oxidants. Ozone is a long-established surrogate for 

31 ambient photochemical oxidants, among which it is by far the most widely studied with regard to 

32 effects on welfare and specifically on vegetation. The information available in this review adds 

33 to our understanding of the atmospheric chemistry for photochemical oxidants and O3 in 
6-2 

1 particular (as described in the ISA, sections 3.2 and 3.6, and summarized in section 2.2 in this 

2 document). The 1996 Staff Paper noted that the database on vegetation effects is generally 

3 considered to raise concern at levels found in the ambient air for O3 and, therefore, control of 

4 ambient O3 levels has previously been concluded to provide the best means of controlling other 

5 photochemical oxidants of potential welfare concern (U.S. EPA, 1996b, p. 277). In the current

6 review, while the complex atmospheric chemistry in which O3 plays a key role has been 

7 highlighted, no alternatives to O3 have been advanced as being a more appropriate surrogate for 

8 ambient photochemical oxidants. Ozone continues to be the only photochemical oxidant other 

9 than nitrogen dioxide that is routinely monitored and for which a comprehensive database exists 

10 (U.S. EPA, 2013, section 3.6). Thus, staff concludes that O3 remains the appropriate pollutant 

11 indicator for use in a secondary NAAQS that provides protection for public welfare from 

12 exposure to all photochemical oxidants. 

13 6.2 FORM AND AVERAGING TIME

14 In considering potential forms and averaging times alternative to that of the current

secondary standard (i.e. 4th 15 high daily maximum 8-hour average, averaged over 3 years) we 

16 address several specific questions.

17 • To what extent does the currently available information provide support for 

18 considering forms different from that of the current secondary standard? 

19 In characterizing the current evidence, the ISA states that “[n]o recent information is 

20 available since 2006 that alters the basic conclusions put forth in the 2006 and 1996 O3 AQCDs” 

21 with regard to biologically relevant exposure indices (U.S. EPA, 2013, section 2.6.6.1). Based 

22 on the current state of knowledge and the best available data assessed in this review, the ISA 

23 therefore concludes that exposure indices that cumulate and differentially weight the higher 

24 hourly average concentrations over a season and also include the mid-level values continue to 

25 offer the most defensible approach for use in developing response functions and in defining 

26 indices for vegetation protection (U.S. EPA, 2013, section 2.6.6.1). In particular, the available 

27 body of evidence provides a wealth of information, compiled over several decades, on the 

28 aspects of O3 exposure that are most important in influencing plant response. As discussed in the 

29 ISA, the importance of the duration of the exposure and the relatively greater importance of 

30 higher concentrations (over lower concentrations) in determining plant response to O3 have been 

31 well documented (U.S. EPA, 2013, section 9.5.3). Building on this research, other work has 

32 been focused on developing “mathematical approaches for summarizing ambient air quality 

33 information in biologically meaningful forms for O3 vegetation effects assessment purposes ...” 

34 (U.S. EPA, 2013, section 9.5.3), including those known as cumulative, concentration weighted 
6-3 

1 forms (i.e. SUM06, W126). Much of this work was completed by the mid-1990s, and was 

2 summarized in the 1996 Criteria Document (CD) (U.S. EPA, 1996a, section 5.5). 

3 On the basis of this longstanding and extensive evidence demonstrating that the risk to 

4 vegetation comes from cumulative seasonal exposures, the EPA in the 1997 and 2008 reviews, as 

5 well as in the 2010 proposed rulemaking to reconsider the 2008 decision, recognized the 

6 importance of cumulative, seasonal exposures as a primary determinant of plant responses to O3

7 in ambient air (61 FR 65741-42; 62 FR 38878; 72 FR 37893, 37896, 37900, 37904; 73 FR 

8 16488-90, 16493-94; 75 FR 3000, 3010, 3012). For example, in the 1996 notice of proposed 

9 rulemaking, the Administrator recognized that the scientific evidence supported the conclusion 

10 that “a cumulative seasonal exposure index is more biologically relevant than a single event or 

11 mean index” (61 FR 65742). In the 2008 review, CASAC recognized that an important 

12 difference between the effects of short-term exposures to O3 on human health and the effects of 

13 O3 exposures on welfare is that “vegetation effects are more dependent on the cumulative 

14 exposure to, and uptake of, O3 over the course of the entire growing season” (Henderson, 2006). 

15 In that review, the CASAC O3 Panel members were unanimous in supporting the final Staff 

16 Paper recommendation that “protection of managed agricultural crops and natural terrestrial

17 ecosystems requires a secondary Ozone NAAQS that is substantially different from the primary 

18 ozone standard in averaging time, level and form” (Henderson, 2007). Accordingly, in both the 

19 1997 and 2008 reviews as well as the 2010 reconsideration, the Administrator proposed a 

20 secondary standard with a cumulative seasonal form as an appropriate policy option (61 FR 

21 65742-44; 72 FR 37899-905; 75 FR 3012-3027). 

22 Over the past two decades, several indices have been considered with regard to their 

23 representation of O3 exposure in terms best describing its effects on vegetation. The 1996 CD 

24 and 1996 Staff Paper evaluated different types of forms to determine which performed the best in 

25 describing the available data on plant response to O3 exposures. These documents noted that a 

26 number of forms (e.g., the one event, mean and unweighted cumulative SUM00) are unable to 

27 reliably predict plant response because they either ignore the role of duration or ignore the 

28 disproportionate impact of higher concentrations by weighting all concentrations equally (U.S. 

29 EPA, 1996b, p. 224). Other forms that were considered include multicomponent forms. 

30 Because these forms take into account many other relevant factors (e.g., plant growth stage, 

31 predisposition from earlier exposures), they consistently predict plant response best of all the 

32 different exposure forms. However, due to being species-specific and highly complex, they were 

33 not considered suitable for more general application in the context of standard setting (U.S. EPA, 

34 1996b, pp. 224-225). On the other hand, it was also found that concentration-weighted forms 

35 that take into account the role of duration and concentration perform almost as well as the 

36 multicomponent forms. These include several threshold forms (e.g., SUM06, AOT60) and 
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1 sigmoidally weighted cumulative indices (e.g., W126) (U.S. EPA, 1996a, pp. 5-84 to 5-136; U.S. 

2 EPA, 1996b, pp. 223-227). Given that these cumulative concentration-weighted forms all were 

3 able to similarly predict plant response on the datasets for which they were evaluated (i.e. 

4 NCLAN), it was not possible to distinguish between them on this basis. Partly as a result, 

5 CASAC deliberations in 1995 did not produce a consensus on which form would be best suited 

6 for a secondary NAAQS. As discussed further in 6.3 below, a workshop held in January of 1996 

7 provided a consensus recommendation on the SUM06 form as appropriate for use in secondary 

8 standards, while also recognizing that a W126 form could also be appropriate (Heck and 

9 Cowling, 1997). Subsequent to this, the final 1996 Staff Paper and 1996 proposal notice both 

10 identified the SUM06 form as appropriate to consider and propose, respectively (U.S. EPA, 

11 1996b, p. 285, 61 FR 65716). In selecting the SUM06 form, it was noted that while it imposed a 

12 threshold even while evidence was lacking for a discernible threshold for O3-related vegetation 

13 effects, it had the benefit of not including concentrations considered to be from background 

14 sources, which was an important feature (U.S. EPA, 1996b, pp. 223-227). 

15 In the subsequent review, the form of the standard was revisited in light of continued 

16 evidence that there remained a lack of discernible threshold for vegetation effects in general, and 

17 newer estimates of O3 concentrations associated with background sources that were lower than in 

18 the previous review so that their inclusion was less of a concern. On these bases, the 2007 Staff 

Paper recommended consideration of the W126 index1 19 as the basis for the form of a distinct 

20 secondary standard (U.S. EPA, 2006, pp. 9-11 to 9-15 and pp. AX9-159 to AX9-187; U.S. EPA, 

21 2007, pp. 7-15/16). The EPA then proposed two options for the secondary standard, one of 

22 which was to adopt a cumulative, seasonal standard based on the W126 index, while the other 

23 option was a secondary standard identical to the proposed revised primary standard (72 FR 

24 37818). The CASAC Panel in that review expressed preference for the W126 index (Henderson, 

25 2006). In EPA’s proposed reconsideration of the 2008 decision, the Agency proposed only the 

26 option of a cumulative, seasonal standard based on the W126 function (75 FR 2938). 

27 In this review, we conclude that specific features associated with the W126 index still 

28 make it the most appropriate cumulative concentration-weighted form for use in the context of 

29 the secondary O3 NAAQS review. In particular, the W126 index does not apply an arbitrary 

30 exposure threshold below which concentrations are not included. Given the acknowledged range 

31 in vegetation sensitivity and the continued lack of evidence of an exposure threshold for effects 

32 above a W126 index of zero, such a feature is desirable. In addition, we conclude the W126 

 

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 

1 sigmoidal weighting function provides the best way to weight concentrations associated with 

2 background sources. Thus, we conclude that the W126 form is best matched to the evidence 

3 associated with vegetation effects, as well as addressing the policy-relevant issue of how to 

4 weight exposures associated with background sources. 

5 • To what extent does the currently available information provide support for 

6 consideration of a cumulative seasonal form derived as a sum of weighted O3

7 concentrations over daylight hours (8:00 am to 8:00 pm) for the 3 month period 

8 having the highest sum within the O3 season? 

9 For a standard with a cumulative seasonal form, including one defined in terms of the 

10 W126 exposure index, the NAAQS element of “averaging time” is appropriately considered in 

11 terms of exposure periods – diurnal and seasonal -- over which the index would be summed in 

12 any given year. In the EPA’s consideration of such a standard in both the 1997 and 2008 

13 reviews, the diurnal period of interest was identified to be the daylight hours, which were defined 

14 for these purposes as spanning the 12-hour period from 8 am to 8 pm during the growing season. 

15 The seasonal period of interest was identified as the consecutive 3 month period with the 

16 maximum W126 index value that occurs within the monitored O3 season. Since many species of 

17 vegetation have growing seasons longer than 3-months in some locations within the U.S., this 

18 seasonal period is viewed as the minimum duration that can appropriately serve as a surrogate 

19 for the range of growing seasons associated with U.S. vegetation (72 FR 37883; U.S. 2013, 

20 section 9.5.3, p. 9-112/13. As discussed below, the evidence available in this review continues 

21 to provide support for use of these periods in considering a cumulative seasonal secondary 

22 standard (U.S. EPA, 2013, section 9.5.3). 

23 For the majority of plants, the diurnal conditions of maximum O3 uptake occur mainly 

24 during the daytime hours (U.S. EPA, 2013, section 9.5.3.2). This is because, in general, (1) 

25 plants have the highest stomatal conductance during the daytime; (2) atmospheric turbulent 

26 mixing is greatest during the day in many areas; (3) the high temperature and high light 

27 conditions that typically promote the formation of tropospheric O3 also promote physiological 

28 activity in vegetation (U.S. EPA, 2013, section 9.5.3.2). In consideration of such a cumulative 

29 metric in the prior review, the EPA has identified the 12-hour period from 8 am to 8 pm as 

30 appropriately capturing the diurnal window with most relevance to the photosynthetic process 

31 (72 FR 37900; 75 FR 3013). In so doing, the EPA recognized, as did CASAC that in some parts 

32 of the country this period may not include all daytime hours or exposures of importance to 

33 vegetation, thus potentially underestimating the impact of O3 at those sites (72 FR 37900-01; 75 

34 FR 3013-14; Henderson, 2007, p. 3, pp. C-22-23). 

35 In addition to daytime uptake, a number of studies have also reported O3 uptake at night 

36 in some species (U.S. EPA, 2013, section 9.5.3.2). Typically the rate of stomatal conductance at 
6-6 

1 night is much lower than during the day. Several field studies have attempted to quantify night2 time O3 uptake with a variety of methods. Across the studies discussed in the ISA, nocturnal 

3 conductance ranged from negligible to 25% of daytime values (U.S. EPA, 2013, section 9.5.3.2). 

4 In some studies the percent of nocturnal uptake varied by season and drought conditions. 

5 However, many of these studies did not link the night-time flux to measured effects on plants, 

6 making it difficult to know whether the impacts on the plant from nocturnal exposures are 

7 greater or less than those from similar daytime exposures, and whether or not they should be 

8 considered as separate impacts or as additive or synergistic with impacts from the preceding 

9 daytime exposure. In addition to the uncertainties associated with understanding the plant 

10 response to nocturnal uptake, there are also uncertainties associated with the extent of its 

11 occurrence. For significant nocturnal stomatal flux and O3 effects to occur, a susceptible plant 

12 with nocturnal stomatal conductance and low defenses must be growing in an area with relatively 

13 high night-time O3 concentrations (often high elevation sites) and appreciable nocturnal 

14 atmospheric turbulence. It is unclear how many areas there are in the U.S. where these conditions 

15 occur. It may be possible that these conditions exist in mountainous areas of southern California, 

16 front-range of Colorado and the Great Smoky Mountains of North Carolina and Tennessee. 

17 More information is needed in locations with high night-time O3 to assess the local O3 patterns, 

18 micrometeorology and responses of potentially vulnerable plant species (U.S. EPA, 2013, 

19 section 9.5.3.2). 

20 In consideration of such uncertainties regarding the importance and extent of nocturnal 

21 exposures associated with plant uptake, and whether and how they might be incorporated into a 

22 national index, staff continues to focus on the 12-hour daytime exposure period of 8:00 am to 

23 8:00 pm, consistent with CASAC advice in the last review (Henderson, 2007, p. 3). In so doing, 

24 we recognize the variability in daylight hours across the country and across the O3 seasons, such 

25 that this period will not always include all the daylight hours in all areas. Thus, a focus on this 

26 set period for all locations has the potential to underestimate the impact of O3 exposures in some 

27 areas, although, given that available monitoring data indicates that the daily increase in O3

28 concentrations generally does not begin until after 8 am (U.S. EPA, 2013, section 3.6.3.2), as 

29 well as our additional focus on the 3-month period with the highest index (as discussed below), 

30 the extent of such underestimation is unclear. We note, however, the variability in daylight 

31 hours nationwide in areas to which a W126-based secondary standard would apply as well as in 

32 areas where studies have been conducted that form the basis for consideration of such a standard 

33 is an important consideration in identifying the appropriate elements for a W126-based 

34 secondary standard. 

35 With regard to identification of the seasonal period over which to cumulate exposures, 

36 EPA notes that a plant is vulnerable to O3 pollution as long as it has foliage and is
6-7 

1 physiologically active (U.S. EPA, 2013, section 9.5.3, p. 9-112), i.e. during its growing season. 

2 The length of vegetative growing seasons varies depending on the type or species of vegetation 

3 and where it grows. For example, as discussed in the ISA, annual crops are typically grown for 

4 periods of two to three months while perennial species may be photosynthetically active longer, 

5 and up to 12 months each year for some species (U.S. EPA 2013, section 9.5.3, p. 9-112). In 

6 general, the period of maximum physiological activity and thus, potential O3 uptake for 

7 vegetation coincides with some or all of the intra-annual period defined as the O3 season, which 

8 varies on a state-by-state basis (U.S. EPA, 2013, Figure 3-24, p. 3-83). This is because the high 

9 temperature and high light conditions that typically promote the formation of tropospheric O3

10 also promote physiological activity in vegetation (U.S. EPA, 2013, section 9.5.3, p. 9-112). 

11 The exposure periods used in studies of O3 effects on vegetation reflect this 

12 understanding, with crop studies typically using shorter seasonal exposure periods, while studies 

13 of longer lived trees and other perennial vegetation often extend for the entire annual growing 

14 season or in some cases over multiple growing seasons. Specifically, the ISA notes that “[m]ost 

15 of the crop studies done through NCLAN had exposures less than three months with an average 

16 of 77 days. Open-top chamber studies of tree seedlings, compiled by the EPA, had an average 

17 exposure of just over three months or 99 days. In more recent FACE experiments, SoyFACE 

18 exposed soybeans for an average of approximately 120 days per year and the Aspen FACE 

19 experiment exposed trees to an average of approximately 145 days per year of elevated O3, 

20 which included the entire growing season at those particular sites. Further, the U.S. Forest 

21 Service and federal land managers have typically used the 6 months from April through 

22 September as the accumulation period (U.S, EPA, 2013, section 9.5.3.2, p. 9-112). However, 

23 despite the possibility that plants may be exposed to ambient O3 longer than 3 months in some 

24 locations, the ISA notes that “[t]he exposure period in the vast majority of O3 exposure studies 

25 conducted in the U.S. has been much shorter than 6 months...” and “there is generally a lack of 

26 exposure experiments conducted for longer than 3 months” (U.S. EPA, 2013, section 9.5.3.2, p. 

27 9-112) 

28 As a result, analyses of effects in terms of the W126 exposure index have typically 

29 defined the index in terms of a 3-month exposure period or at least in terms of periods shorter 

30 than 6 months (e.g. SoyFACE, AspenFACE) (U.S, EPA, 2013, p. 112). In addition, the O3

31 season within which O3 monitoring is required is shorter than 6 months in many areas in the 

32 country, yet longer than 3 months in all locations (U.S. EPA, 2013, section 3.6.2, p. 3-83). Thus, 

33 in the last review, the EPA proposed use of the consecutive 3-month period within the O3 season

34 having the maximum W126 index value. A 3-month exposure period was also supported by 

35 CASAC in advice provided both during the last review and on the 2010 proposed reconsideration 

36 (Henderson, 2006; Samet, 2010). In revisiting this issue in the current review, the EPA 
6-8 

1 conducted a new analysis to further inform the consideration of the most appropriate seasonal 

2 accumulation period (U.S. EPA, 2013, section 9.5.3). This analysis calculated and compared the 

3 3- and 6-month maximum W126 values for over 1,200 AQS and CASTNET EPA monitoring 

4 sites for the years 2008-2009. The two accumulation periods were found to have highly 

5 correlated metrics (U.S. EPA, 2013, Figure 9-13). The two accumulation periods were centered 

6 on the yearly maximum for each monitoring site, and it is possible that this correlation would be 

7 weaker if the two periods were not temporally aligned (U.S. EPA, 2013, section 9.5.3). The 

8 analysis indicates that in the U.S., W126 cumulated over 3 months and W126 cumulated over 6 

9 months are proxies of one another, as long as the period in which daily W126 is accumulated 

10 corresponds to the seasonal maximum. Therefore, it is expected that either statistic will predict 

11 vegetation response equally well. 

12 • Are there other aspects of the form that are appropriate to consider that 

13 contribute to welfare protection provided by a cumulative seasonal standard? 

14 In considering a secondary standard in terms of the W126 exposure index with daylight 

15 12-hour averaging time and maximum consecutive 3-month cumulative exposure period, we find 

16 it appropriate to evaluate the protection that might be afforded by a form limited to a single year 

17 or one that is based on evaluation of W126 index values across multiple years. In so doing, we 

18 are mindful that the protection provided by the secondary standard derives from the combination 

19 of all elements of the standard. Accordingly, the discussion below explores the information 

20 relevant to consider in conjunction with identification of standard level, as well as whether there 

21 is a lack of support in the current information for either single or multi-year options. 

22 In considering an annual form of a standard, we particularly take note of O3-induced 

23 vegetation effects that can occur as a result of a single year’s exposure. These include visible 

24 foliar injury symptoms, growth reduction in annual species, and crop yield loss in annual crops. 

25 The following discussion considers these effects, in the context of their potential public welfare 

26 significance, with regard to the extent to which a W126-based standard with an annual form or 

27 one based on evaluation across multiple years may be able to provide appropriate protection. 

28 Visible foliar injury that occurs on public lands that have been afforded special 

29 protections (e.g., Class I areas) is an annual effect with potential public welfare significance, 

30 particularly when the injury is of some severity. However, we recognize the complexity 

31 associated with defining O3 exposure conditions that might consistently provide appropriate 

32 protection from this endpoint. As summarized in section 5.4.2 and described more completely in 

33 the WREA, over 80% of USFS biosites show no foliar injury in a wide variety of tree species 

34 under air quality conditions (2006-2010) that include a broad range of W126 index values (U.S. 

35 EPA 2014, section 7.2). The WREA USFS biosite analysis additionally suggested a reduction in 

36 prevalence of sites with visible foliar injury with W126 index values below approximately 10 
6-9 

1 ppm-hrs. However, the full body of evidence indicates wide variability in this endpoint, such 

2 that although evidence shows visible foliar injury can occur under very low cumulative O3

3 concentrations, “...the degree and extent of visible foliar injury development varies from year to 

4 year and site to site..., even among co-members of a population exposed to similar O3 levels, due 

5 to the influence of co-occurring environmental and genetic factors” (U.S. EPA 2013, section 

6 9.4.2, p. 9-38). In particular, drought conditions which can frequently occur with high seasonal 

7 O3 concentrations can potentially offer some protection so that the visible foliar injury that might 

8 be expected based only on O3 exposure conditions does not develop. Accordingly, staff 

9 recognizes the challenges associated with characterizing the role of relative soil moisture, and its 

10 appreciable spatial and temporal variability. For example, the WREA analysis of the 

11 relationship between foliar injury at USFS biosites and soil moisture data found foliar injury to 

12 be less prevalent in years when sites were drier (U.S. EPA, 2014, section 7.2 and Figure 7-10). 

13 The uncertainties associated with describing the potential for the occurrence of foliar injury and 

14 its severity or extent of occurrence for any given set of O3 exposure conditions, limit our ability 

15 to identify annual ambient O3 exposure conditions that might be expected to provide an 

16 appropriate degree of protection. Thus, the evidence for this endpoint does not appear to provide 

17 support for the necessity of consideration of a single year form. 

18 Another annual effect with the potential for public welfare significance is that of reduced 

19 yields in annual commodity crops. It has been well documented that certain important 

20 commodity crops are sensitive to O3 and that achieving the optimum level of crop yields is of 

21 special importance to the public welfare. Although the impact of this effect can be monetized, 

22 determining what the optimum yields are is less straightforward. This is because optimizing the 

23 public welfare associated with annual crop production is not necessarily the same as maximizing 

24 the annual yield but instead by achieving the maximum levels of producer and consumer 

25 surpluses. These latter objectives depend less on the magnitude of the annual O3 exposure and 

26 more on the management strategies employed by the agricultural community during that year, 

27 strategic planting choices that are made from year to year, and environmental factors (e.g., 

28 drought) which can in large part offset these concerns (U.S. EPA, 2014, Chapter 6, section 6.5). 

29 Thus, since the optimum annual yield is not necessarily equivalent to the greatest public welfare 

30 value, it is difficult to know what degree of protection from O3 would be requisite in any given 

31 year. Thus, as in the 2008 review, we do not believe that a public welfare benefit of optimizing 

32 crop yields by protection of commodity crops from annual O3 impacts is necessarily achievable 

33 through the secondary NAAQS (73 FR 16492, 16497). Further, depending on the level of 

34 protection being considered for other vegetation effects (i.e. seedling biomass loss), crops may 

35 indirectly receive protection. Thus, similar to the annual foliar injury effect, the uncertainties 

36 associated with identifying at what point annual yield loss in commodity crops becomes adverse 
6-10 

1 to the public welfare make it difficult to identify an appropriate degree of protection, particularly 

2 when considering the effect of a single year. 

3 In contrast to impacts on annual species that accrue in the single growing season in which 

4 the O3 exposures occur, annual effects in perennial species can be “carried over” into the 

5 subsequent year where they affect growth and reproduction (U.S. EPA, 2013, pp. 9-43 to 9-44 

6 and p. 9-86). In addition, when these effects occur over multiple years due to elevated O3

7 exposures across several years, they are compounded, increasing the potential for effects at the 

8 ecosystem level and associated ecosystem services that may be of significance to the public 

9 welfare. For example, in perennial species, O3-induced reduction in photosynthesis decreases 

10 the amount of carbohydrate produced by and available to the plant for growth and storage. The 

11 plant often responds to decreases in carbohydrate production by reallocating the remaining 

12 carbohydrate to above ground growth at the expense of storage in and growth of the roots (U.S. 

13 EPA, 2013, section 9.4.3.1). When fewer carbohydrates are available in the roots, less energy is 

14 available for root-related functions such as water and nutrient acquisition. By inhibiting the 

15 amount of carbohydrate transfer to the roots, O3 also can disrupt the symbiotic relationship that 

16 exists between some plants and mycorrhizal fungi in the soil. Mycorrhizal fungi have been 

17 shown to improve the uptake of nutrients, protect host plant roots against pathogens, and produce 

18 plant growth hormones (U.S. EPA, 2007, 7.3.3.2). Thus, such effects of elevated O3 in a given 

19 year may affect the plant’s resistance in a subsequent year, with multiple such years having the 

20 potential for larger impacts. However, where there is variation across years such that lower O3

21 years with lesser impact occur between elevated O3 years, it may be that the longer-term impacts 

22 are tempered. 

23 Effects of elevated O3 years on perennial plants, when they occur over several years, may 

24 also contribute to effects on ecosystem services, i.e. alteration of below-ground biogeochemical 

25 cycles, and alteration of both above- and below- ground terrestrial community composition and 

26 terrestrial ecosystem water cycling (U.S. EPA, 2013, Table 9-19). Ozone has also been shown to 

27 affect plant reproduction in numerous ways (U.S. EPA, 2007, 7.3.3.3; U.S. EPA, 2013, 9.4.3.1). 

28 These effects, when they occur at sufficient magnitude for a single species, may result in 

29 impaired recruitment and loss of the species from the stand or community. This has the potential 

30 to change the community composition and biodiversity. If these effects occur in multiple plant 

31 species and/or over multiple years, they can result in a reduction in the productivity and carbon 

32 sequestration of terrestrial ecosystems. Such ecosystem-related effects and others discussed in 

33 the ISA may be considered to reflect impacts of critical O3 exposures over the longer term. We 

34 additionally note that as compared to intermittent (or single year) critical O3 exposures, multiple 

35 years of such exposures might be expected to result in larger impacts on forested areas, i.e. 
6-11 

1 increased susceptibility to other stressors such as insect pests, disease, co-occurring pollutants 

2 and harsh weather, due to the compounding or carry-over effects on tree growth. 

3 Given the above, we find it reasonable to conclude that the public welfare significance of 

4 the effects that can occur as a result of multiple-year O3 exposures are greater than those 

5 associated with a single year. Thus, to the extent that the Administrator’s focus for public 

6 welfare protection to be afforded by the secondary O3 standard is on long-term effects that occur 

7 in sensitive tree species in natural forested ecosystems including in federally protected areas such 

8 as Class I areas or on lands set aside by States, Tribes and public interest groups to provide 

9 similar benefits to the public welfare, a standard with a form that evaluates the cumulative 

10 seasonal index across multiple years (in combination with an appropriate level) might be 

11 considered to provide an appropriate match to the nature of O3-related effects on vegetation most 

12 at risk of being adverse to the public welfare and thus, upon which the secondary O3 standard is 

13 focused. 

14 Although cumulative, seasonal exposure values of interest for vegetation effects have 

15 been expressed in terms of a single season (e.g., W126 index values for a season’s exposure of 

16 tree seedlings), we recognize that it is also appropriate to consider a form that is evaluated over a 

17 multiple-year period, such as three years (U.S. EPA, 2007; 72 FR 37901; 75 FR 3021). Such an 

18 aspect to the form specifies the number of years for which monitoring data are considered in 

19 judging attainment with a standard. For example, under a standard with a single year form, a 

20 monitor may be judged to meet the standard based on a single year of data, while under a 

21 standard with a form requiring evaluation over a multi-year period, a monitor is not judged to 

22 have met the standard until a complete multi-year record is available. For metrics with 

23 substantial year-to-year variability, a multi-year evaluation period can provide stability in air 

24 quality management programs, thus facilitating achievement of the protection intended by a 

25 standard, while a standard with an annual form has the potential to contribute to challenges to 

26 implementation of air pollution control programs in areas identified as not meeting the standard 

27 that, through variability in the metric, meet the standard level in other years. 

28 For a W126-based potential standard, the multi-year form identified for consideration in 

29 the last review was the average cumulative seasonal metric over three consecutive years (75 FR 

30 3027). Such a multi-year form remains appropriate to consider to provide stability to an 

31 alternative secondary standard, just as the multi-year form provides for the current standard 

32 (average over three years of annual fourth-highest daily maximum 8-hour average O3

concentrations).2 33 In considering the issue of stability in the context of such a form, we first note 

34 the inter-annual variability of seasonal W126 index, which is not unexpected given the logistic 

 

2 See ATA III, 283 F. 3d at 374-75 (recognizing programmatic stability as a legitimate consideration in the 

NAAQS standard-setting process).
6-12 

1 weighting function and also inter-annual variability in meteorological conditions which 

2 contribute to O3 formation (see Appendix 2C). The staff analysis in Appendix 2C describes the 

3 variability in annual W126 values in relation to variability in the 3-year average, which indicates 

4 that a standard based on an annual W126 index would be expected to have a lower degree of 

5 year-to-year stability relative to a standard based on a form that averages seasonal indices across 

6 three consecutive years. A more stable standard can be expected to contribute to greater public 

7 welfare protection by limiting year-to-year disruptions in ongoing control programs that would 

8 occur if an area was frequently shifting in and out of attainment due to extreme year-to-year 

9 variations in meteorological conditions. 283 F. 3d at 374. Thus, we recognize the public welfare 

10 benefit of having a standard of a form with more year-to-year stability. 

11 In considering the relative value of a multi-year form for an alternative secondary 

12 standard in affording the desired protection, we also take note of the influence such a form might 

13 be expected to have on air quality conditions and the associated public welfare protection. For 

14 example a W126-based standard with an annual form would have a level defining the maximum 

15 cumulative seasonal index. Where air quality is attained that reliably meets such a standard, the 

16 cumulative seasonal index values would be at or lower than the level of the standard in all years 

17 and, given the significant inter-annual variability in seasonal W126 values (described above), 

18 would be appreciably below the standard level in many years. Alternatively, for a standard with 

19 a form that averages the cumulative seasonal index values across three consecutive years, the 

20 seasonal index would be above the level in some years, but would have to be below it in others 

21 within the same 3-year period, restricting there to no more than two years out of three to have 

22 indices above the level, and depending on magnitude of each year’s index, potentially no more 

23 than one. 

24 In our consideration of this aspect of form for a cumulative seasonal secondary standard, 

25 we additionally take note of advice from CASAC on this topic in prior reviews. For example, in 

26 comments provided on the final Staff Paper in 2007, CASAC expressed the view that “[i]f multi27 year averaging is employed to increase the stability of the secondary standard, the level of the 

28 standard should be revised downward to assure that the desired threshold is not exceeded in 

29 individual years” (Henderson, 2007). Accordingly, in considering all elements for a revised 

30 standard, including level, as well as form, we note the importance of considering the potential for 

31 single-year O3 exposures that would result in adverse effects to public welfare. As noted above, 

32 a standard with a form that averages across three years can also control year-to-year variability 

33 and individual year concentrations. In so noting, we recognize the availability of air quality 

34 analyses that can inform our consideration of the likely extent of such control (e.g., Appendix 

35 2C). Additionally, we note the potential that, depending on type of effects and the magnitude 

36 which may be judged adverse, low O3 years within each three-year period might play a 
6-13 

1 countervailing role with regard to the extent of O3-attributable impacts that might pose public 

2 welfare significance. The WREA analyses on the potential for biomass loss estimates based on a 

3 3-year average to underestimate the cumulative impact on growth inform this point. Those 

4 limited analyses indicate the potential for underestimation, although of relatively small 

5 magnitude (U.S. EPA, 2014, section 6.2.1.4). We further recognize the role of public policy 

6 judgments in drawing conclusions regarding adversity of effects to public welfare.

7 In light of the relationships described above, the appropriate level and form combination 

8 will depend on the Administrator’s objectives for the target level of protection. In articulating 

9 these objectives it will be appropriate to evaluate the nature of the O3 induced effects and their 

10 significance or importance to the public welfare, as well as the role that year-to-year exposure 

11 variability can play in public welfare impacts. For example, to the extent that the 

12 Administrator’s priority for public welfare protection is against long-term effects in natural 

13 forested ecosystems, and given uncertainty with regard to the extent to which effects from 

14 single-year exposures may be judged adverse, articulation of a target level of protection in terms 

15 of a longer-term condition, such as through a form that averages the index across multiple years, 

16 may be appropriate. 

17 • What does the available information indicate with regard to protection of welfare 

18 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 19 high 

20 8-hour average concentrations)?

21 In staff consideration of the primary standard in chapter 4, staff preliminarily concludes it 

22 is appropriate to consider alternative primary standards of the same form and averaging time as 

23 the current primary standard. Thus, although the discussion in this chapter, with regard to the 

24 secondary standard, indicates the appropriateness of considering an alternative secondary 

25 standard with a cumulative seasonal form, we also recognize that, to the extent that the 

26 Administrator may find it more effective to control air quality using the same form for both the 

27 primary and secondary standards, it may be practical to consider the extent to which a standard 

28 in the form of the primary standard might be expected to also reduce and provide protection from 

29 cumulative seasonal exposures of concern. For example, if a clear and robust relationship was 

30 found to exist between 8-hour daily peak O3 concentrations and cumulative, seasonal exposures, 

31 the averaging time and form of the current standard might be concluded to have the potential to 

32 be effective as a surrogate. Addressing this point, the ISA describes the results of a recent focus 

33 study that examined the diel variability in O3 concentrations in six rural areas between 2007 and 

34 2009 (U.S. EPA, 2013, pp. 3-131 to 3-133). The ISA reported that “[t]here was considerable 

35 variability in the diel patterns observed in the six rural focus areas” with the three mountainous 

36 eastern sites exhibiting a “generally flat profile with little hourly variability in the median 
6-14 

1 concentration and the upper percentiles”, while the three western rural areas demonstrated a 

2 “clear diel pattern to the hourly O3 data with a peak in concentration in the afternoon similar to 

3 those seen in the urban areas”, which was especially obvious at the San Bernardino National 

4 Forest site, 90 km east of Los Angeles at an elevation of 1,384 meters (U.S. EPA, 2013, p. 3-

5 132). Thus, while the western sites that are influenced by upwind urban plumes may have 

6 increased cumulative seasonal values coincident with increased daily 8-hour peak O3

7 concentrations, this analysis indicates that, in sites without such an urban influence (the eastern 

8 sites in this analysis), such a relationship does not occur (U.S. EPA, 2013, section 3.6.3.2). 

9 Thus, the lack of such a relationship indicates that in some locations, O3 air quality patterns can 

10 lead to elevated cumulative, seasonal O3 exposures without the occurrence of elevated daily 

11 maximum 8-hour average O3 concentrations (U.S. EPA, 2013, section 3.6.3.2). Further staff 

12 notes that the prevalence and geographic extent of such locations is unclear, since as in the last 

13 review, there continue to be relatively fewer monitors in the West, including in high elevation 

14 remote sites. In considering the findings of this analysis, we additionally recognize, however, 

15 that in general, the cumulative seasonal values for the eastern rural sites, where cumulative 

16 seasonal O3 concentrations appear to be relatively less related to daily maximum 8-hour 

17 concentrations, are lower than those of the western, urban-influenced sites.

18 In addition to the focus study described in the ISA (U.S. EPA, 2013, section 3.6.3.2), we 

19 considered analyses of air quality monitoring data and air quality modeling analyses. Chapter 2 

20 of this document characterizes recent monitoring data on O3 air quality in rural areas. While 

21 approximately 80% of the O3 monitoring network is urban focused, about 120 rural monitors are 

22 divided among CASTNET, NCore, and POMs sites (Chapter 2, pp. 2-2 to 2-3, Figure 2.1). 

23 Specifically, as stated in chapter 2 “[a]lthough rural monitoring sites tend to be less directly 

24 affected by anthropogenic pollution sources than urban sites, rural sites can be affected by 

25 transport of O3 or O3 precursors from upwind urban areas and by local anthropogenic sources 

26 such as motor vehicles, power generation, biomass combustion, or oil and gas operations” (U.S. 

27 EPA, 2013, section 3.6.2.2). In addition, O3 tends to persist longer in rural than in urban areas 

28 due to lower rates of chemical scavenging in non-urban environments. At higher elevations, 

29 increased O3 concentrations can also result from stratospheric intrusions (U.S. EPA, 2013, 

30 sections 3.4, 3.6.2.2). As a result, O3 concentrations measured in some rural sites can be higher 

31 than those measured in nearby urban areas (U.S. EPA, 2013, section 3.6.2.2) and the ISA 

32 concludes that “cumulative exposures for humans and vegetation in rural areas can be 

33 substantial, and often higher than cumulative exposures in urban areas” (U.S. EPA, 2013, p. 3-

34 120). These known differences between urban and rural sites suggest that there is the potential 

35 for an inconsistent relationship between 8-hour daily peak O3 concentrations and cumulative, 
6-15 

1 seasonal exposures in those areas where protection of O3-sensitive vegetation may be most 

2 needed.

3 In addition, as was done in both the 1997 and 2008 reviews, staff has analyzed 

4 relationships between O3 levels in terms of the current averaging time and form and a W126 

5 cumulative form, based on recent air quality data. One analysis describes the W126 index values 

6 and current standard design values at each monitor for two periods: 2001-2003 and 2009-2011 

7 (e.g., Appendix 2B, Figures 2B-2 and 2B-3). This shows that between the two periods, during 

8 which broad scale O3 precursor emission reductions occurred, O3 concentrations in terms of both 

9 metrics were reduced. There is a fairly strong, positive degree of correlation between the two 

metrics (Appendix 2B).3 10 Focusing only on the latter dataset (2009-2011), it can be seen that at 

11 monitors just meeting the current standard (3-year average fourth-highest daily maximum 8-hour 

12 average concentration equal to 0.075 ppm), W126 values (in this case 3-year averages) varied 

13 from less than 9 ppm-hrs to approximately 20 ppm-hrs (Appendix 2B, Figure 2B-3b). At sites 

14 with a 3-year average fourth-highest daily maximum 8-hour average concentration at or below a 

15 potential alternative primary standard level of 70 ppb, 3-year W126 index values range from 5 to 

16 18 ppm-hrs, with very few above 15 ppm-hrs. An alternative presentation of such data for recent 

17 3-year periods back to 2006 – 2008 indicate that among the counties with O3 concentrations that 

18 met the current standard, the number with 3-year W126 index values above 15 ppm-hrs ranges 

19 from fewer than 10 to 24 (Appendix 2B, Figure 2B-9). In general during this longer period, 

20 W126 index values above 15 ppm-hrs and meeting the current standard were pre-dominantly in 

21 Southwest region. As the first analysis in Appendix 2B (for the 2001-2003 and 2009-2011 

22 periods) indicates, monitors in the West and Southwest tend to have higher W126 values relative 

23 to their design values than do monitors in other regions. This pattern is noteworthy because the 

24 Southwest region has a less dense monitoring network than regions in the Eastern US (see Figure 

25 2-1), so that the extent to which this pattern occurs throughout these regions is uncertain. 

26 Although single-year W126 index values were not separately analyzed in this analysis of the 

27 monitor data, it indicates appreciable variation in cumulative, seasonal O3 concentrations among 

28 monitor locations meeting different levels of a standard of the current form. 

29 Analyses of the WREA air quality scenarios indicate the potential for O3 precursor 

30 emission reductions achieving O3 concentrations that just meet different 8-hour standards to

31 produce a significant reduction in 3-year W126 index values. For example, for the current 

32 standard scenario, nearly all model-adjusted monitors are at or below an estimated 3-year 

33 average W126 value of 15 ppm-hrs (as summarized in section 5.2.2 above and described in U.S. 

 

3 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 values.
6-16 

1 EPA, 2014, 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 2 high 8-

3 hour average scenario of 70 ppb, averaged across 3 years, nearly all monitors in the U.S. would 

4 meet a 3-year W126 of 11 ppm-hrs, though some monitors in the southwest would remain 

5 between 11 and 15 ppm-hrs. At 65 ppb, all locations are at or below 11 ppm-hrs. Thus, similar 

6 to the monitoring analysis, the modeling analysis generally indicates reductions in W126 levels 

7 with reduced O3 concentrations in terms of the current standard averaging time and form. This 

8 suggests that depending on the level for a standard of the current averaging time and form, a 

9 degree of welfare protection may be afforded. The extent to which such protection provides 

10 adequate public welfare protection additionally depends on the level of protection identified by 

11 the Administrator for the public welfare in terms of the W126 index. In so noting, however, we 

12 recognize the importance of also considering uncertainties in both the model-adjustment analyses 

13 and those based on monitoring data. These uncertainties, including those related to monitor 

14 coverage, the extent to which recent data can be expected to describe future relationships, and 

15 modeling approaches, among others, should be kept in mind when assessing the strength of this 

16 apparent relationship. 

17 6.3 LEVEL

18 In considering potential levels for alternative secondary standards, we have taken into 

19 account scientific evidence characterized in the ISA and exposure and risk estimates for different 

20 air quality scenarios analyzed in the WREA, as well as the uncertainties and limitations in this 

21 information. We consider this information together with regard to support for alternative 

22 standards that might be appropriate to consider to provide the requisite protection from adverse 

23 effects to public welfare. Drawing from section 6.2 above, our discussion here is primarily in 

24 terms of the W126 sigmoidally-weighted metric, cumulated over the 12 daylight hour period 

25 (8:00 am to 8:00 pm) for the consecutive three month period with the maximum index value 

26 within the growing season. In addition to considering the information in the context of a single 

27 growing season, we also consider it in the context of a form for this W126 metric averaged 

28 across 3 consecutive growing seasons for reasons discussed in section 6.2 above. 

29 In the discussion below, we turn first to consideration of the currently available scientific 

30 evidence as assessed and characterized in the ISA. We then consider the WREA findings with 

31 regard to vegetation, ecosystem and ecosystem services effects or risks estimated for different air 

32 quality scenarios. We additionally take note of important uncertainties and limitations in the 

33 evidence and exposure/risk analyses. Lastly, we take note of judgments to be made by the 

34 Administrator in drawing conclusions regarding effects and risks that represent adverse 

35 environmental effects to public welfare. In so doing, we identify key considerations with regard 
6-17 

1 to the currently available evidence, exposure/risk information and associated uncertainties in 

2 identifying levels that may be appropriate to consider for a cumulative seasonal secondary 

3 standard. Such levels are described in section 6.5 below which describes preliminary staff 

4 conclusions regarding alternative secondary standards appropriate to consider in this review. 

5 • What does the currently available evidence indicate with regard to the range of 

6 W126-based index values that may provide protection from vegetation effects of 

7 O3? 

8 In considering this question in terms of a cumulative seasonal W126-based index, we 

9 consider first quantitative evidence for O3 exposure effects on plant growth, productivity and 

10 related endpoints. In so doing, we draw on the C-R functions derived from the evidence on tree 

11 seedlings and crops assessed in the ISA which are further described in the WREA and 

12 summarized in sections 5.2.1 and 5.3.1 (U.S. EPA 2013, section 9.6; U.S. EPA, 2014, section 

13 6.2). It is important to note that these functions are used to provide estimates of growth and yield 

14 reduction in tree seedlings and crops that might be expected to result from exposure over a single 

15 growing season to various O3 concentrations expressed in terms of a W126 index (Figure 5-1 

16 above). 

17 Table 6-1 below presents estimates of relative yield and biomass loss for the studied 

18 species of crops and tree seedlings, respectively, for a growing season exposure to a number of 

19 W126 levels. In this table, we have included observations related to 2% relative biomass loss in 

20 tree seedlings and 5% relative crop yield loss. These values are consistent with advice from 

21 CASAC thus far in this review (Frey and Samet, 2012) on factors on which to base consideration 

22 of levels for a secondary standard (see section 6.4 below), and with values given focus in the 

23 1996 consensus-building workshop, described below. The observations in Table 6-1 for each 

24 W126 index value include relative biomass or crop yield loss estimates for the median across the 

25 studied species, as well as the proportion of species with estimates below 2%, 5% or 10%, as 

26 examples of benchmarks that may be of interest in different contexts. 

27 Staff recognizes Table 6-1 as a useful way to consider the magnitude of biomass and crop 

28 yield loss estimates across the 12 and 10 studied species, respectively, for the purposes of 

29 informing consideration of potential levels for alternative secondary standards based on the 

30 W126 index. The different index levels can be considered with regard to the magnitude of 

31 median species response and the proportion of species with responses expected to be below the 

32 various benchmarks. For example, at an O3 exposure equivalent to a W126 index value of 7 

33 ppm-hrs, the median tree seedling response across the 12 studied species is projected to be near 

34 or below 2% reduction in growth compared to that for the control plants exposed to W126 index 

35 values of zero (Table 6-1; U.S. EPA, 2014, section 6.2, Appendix 6A). Seven of the 12 species 

36 are projected to have smaller than 2% reduction in growth, while three other species are 
6-18 

1 projected to have a reduction between 2 and 5%, for a total of 10 of 12 species projected to have 

2 less than 5% reduction in growth, at a W126 index value of 7 ppm-hrs. The remaining two 

3 species, black cherry and cottonwood, are projected to have reductions just over 15% and 40%, 

4 respectively, at 7 ppm-hrs and above (Table 6-1). The median growth reduction response 

5 projected for a W126 index value of 15 ppm-hrs is approximately 5%, with responses less than 

6 2% projected for five of the twelve studied species and less than 5% for six, while less than 10% 

7 reduction in growth is projected for ten of the twelve species. Somewhat similar estimates are 

8 projected for a W126 index value of 17 ppm-hrs, with the median growth reduction response 

9 projected to be 6%, and also with five of the twelve species less than 2%, nine less than 10% and 

10 a tenth just over 10% (Table 6-1). Crop yield loss estimates are also described in Table 6-1.

11 
6-19 

1 Table 6-1. Seedling and crop growth or yield reductions estimated for O3 exposure over a 

2 season. 

W126 value

for exposure 

period

Tree seedling growth reduction (biomass loss)

A Crop yield lossA

21 ppm-hrs BMedian species w. 7.7 % loss (varying lower)

3/12 species w. < ~2%C loss; 5/12 w. < 5% loss

7/12 species w. <10% loss; 10/12 <15% loss

Note: Most sensitive 

species are 

cottonwood (relative 

biomass loss 

estimates > 40% for 

W126>7) and black 

cherry (relative 

biomass loss 

estimates >15% for 

W126 > 7).

Median speciesD = 7.7 % loss

4/10 species w. < 5% loss

3/10 species w. >5,<10% loss

3/10 species w. >10,<20% lossE

19 ppm-hrs BMedian species w. 6.8 % loss (varying lower)

3/12 species w. < ~2% loss; 5/12 w. <5% loss

7/12 w. <10% loss; 10/12 w. <15% loss

Median speciesD = 6.4 % loss

5/10 species w. < 5% loss

3/10 species w. >5, <10% loss

2/10 species w. >10,<20% loss

17 ppm-hrs BMedian species w. 6.0 % loss (varying lower)

5/12 species w. < ~2% loss; 5/12 w. <5% loss

9/12 w. <10% loss; 10/12 w. <15% loss

Median speciesD = 5.1 % loss

5/10 species w. < 5% loss

3/10 species w. >5, <10% loss

2/10 species w. >10,<20% loss 

15 ppm-hrs BMedian species w. 5.2 % loss (varying lower)

5/12 species w. < ~2% loss; 6/12 w. <5% loss

10/12 species w. <10% loss 

Median speciesD <5% loss

6/10 species w. < 5% loss

4/10 species w. >5, <10% loss F

13 ppm-hrs BMedian species w. 4.4% loss (varying lower)

5/12 species w. < ~2% loss; 7/12 w. <5% loss

10/12 species w.<10% loss 

Median speciesD <5% loss 

7/10 species w. < 5% loss

3/10 species w. >5, <10% loss

11 ppm-hrs BMedian species w. 3.5% loss (varying lower)

5/12 species w. < ~2% loss; 8/12 w. <5% loss

10/12 species w. <10% loss 

Median speciesD <5% loss 

9/10 species w. < 5% loss

1/10 species w. >5, <10% loss g

9 ppm-hrs BMedian species w. 2.5% loss (varying lower)

5/12 species w. < ~2% loss; 10/12 w. <5% loss

Median speciesD <5% loss 

All species < 5% loss

7 ppm-hrs BMedian species w. <2% loss (varying lower)

7/12 species w. < ~2% loss; 10/12 w. <5% loss

Median speciesD <5% loss 

All species < 5% loss

A Estimates here are based on the C-R functions described in WREA, section 6.2 and discussed in section 5.2.1

B This median value is the composite median from WREA, Figure 6-4 (also discussed in section 5.2.1). The note in parentheses 

refers to the median of the stochastic sampling estimate shown in WREA Figure 6-4 (and Figure 5-2 in this document).

C In making comparisons, biomass loss estimates described as < 2 refer to values < 2.04.

D This median is the composite median from WREA, Figure 6-5 (also discussed in section 5.2.1). 

E Kidney bean , potato, and winter wheat >15%. F Cotton, kidney bean, potato, & soybean >5%. G Kidney bean at 5.7% loss.

3 

4 With respect to W126 index values associated with different degrees of protection for 

5 visible foliar injury, we first recognize the value of this endpoint in its long-standing use as a 

6 well established bioindicator of O3 exposure, as described in the ISA (U.S. EPA 2013, section 

7 9.4.2). 
6-20 

1 In addition to the role of visible foliar injury as an indicator, we note that the aesthetic

2 aspects of visible foliar injury itself have the potential to be important to public welfare (as 

3 described in section 5.4 above). Analyses such as those presented in the WREA, and 

4 summarized in section 5.4.2 above, provide approaches by which foliar injury observations 

5 under existing air quality may be analyzed. However, there is very little information on which to 

6 base judgments of the prevalence and/or severity of injury that might be considered adverse. We 

7 further recognize that there is substantial variability such that “the degree and extent of visible 

8 foliar injury development varies from year to year and site to site ... even among co-members of 

9 a population exposed to similar O3 levels, due to the influence of co-occurring environmental 

10 and genetic factors” (U.S. EPA, 2013, p. 9-38). Thus, based on consideration of the evidence, 

11 staff recognizes the lack of a consistent or generally predictable relationship between particular

12 W126 exposures and visible foliar injury incidence, given the influence of and variability in 

13 relative soil moisture and other environmental factors. We additionally note uncertainty in what 

14 can be concluded from foliar injury in relation to plant health, productivity and ecological 

15 function as “it is not presently possible to determine, with consistency across species and 

16 environments, what degree of injury at the leaf level has significance to the vigor of the whole 

17 plant” (U.S. EPA, 2013, p. 9-39). Thus, on the basis of the evidence, we are not able to identify 

18 a range of appropriate W126 index values. 

19 In further considering the available information pertaining to the question above, we 

20 additionally recognize conclusions that have been drawn by expert committees with regard to 

21 these endpoints (i.e., tree seedling growth, crop yields and visible foliar injury). For example, in 

22 their review of staff documents during the O3 NAAQS review completed in 1997 review, the 

23 CASAC O3 panel members expressed a wide range of opinions on aspects of the evidence 

24 important to consider in judging the adequacy of the O3 secondary standard and in considering 

25 the form and level that would be appropriate for a secondary O3 standard (Wolff, 1996). 

26 Subsequent to CASAC meetings in 1995 on this topic, a consensus-building workshop sponsored 

27 by the Southern Oxidant Study group was held on the topic of the O3 secondary standard in 

28 January 1996 (Heck and Cowling, 1997). This workshop was attended by 16 scientists with 

29 backgrounds in agricultural, managed forest, natural systems, and air quality, all of whom were 

30 leaders in their fields and whose research formed the basis of much of the research examined in 

31 the 1996 Criteria Document. These scientists expressed their judgments on what standard 

32 level(s) would provide vegetation with protection from O3-related adverse effects that would be 

adequate, in their view.4,5 33 As the 1997 workshop publication indicates, the scientists at the 1996 

 

4 At the time of the workshop, the secondary O3 standard being reviewed by EPA was a 1-hour average of 

0.08 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). 
6-21 

1 workshop also reached consensus views regarding the types of exposures that were important in 

2 eliciting plant response and the types of metrics that were best at predicting these responses 

3 (Heck and Cowling, 1997). Before coming to agreement on daily and seasonal durations and 

4 forms pertinent to a distinct secondary standard, the participants discussed and identified 

endpoints to consider for natural, forest and agricultural ecosystems.6 5 With regard to form of the 

6 standard, participants concurred with either the SUM06 or W126 metrics, with consensus finally 

7 reached for SUM06, with some qualification regarding implications for a threshold. The 

8 participants identified the ranges they felt should be considered for each of three endpoints. 

9 Overall, the SUM06 values ranged from 8 to 20 ppm-hr (corresponding to W126 values ranging 

from 5 to 17 ppm-hr, based on EPA analysis focused on conditions in NCLAN studies).7 10 This 

11 overall range reflected ranges for each of the three endpoints, with the following considerations 

12 (Heck and Cowling, 1997). 

13 – Crops (yield reductions): SUM06 of 15-20 ppm-hrs (13-17 ppm-hrs, W126). This 

14 range was recognized to generally consider <10% yield loss in more than 75% of 

15 species.

16 – Trees (growth effects): SUM06 of 10-16 ppm-hrs (7 to 14 ppm-hrs, W126). This 

17 range was recognized to generally consider 1-2% per year growth reduction; in so 

18 doing,the group identified a need to consider the potential for year-to-year 

19 compounding of impacts in long-lived perennial species.

20 – Visible Foliar Injury: SUM06 of 8 to 12 ppm-hrs (5 to 9 ppm-hrs, W126). 

21 Since the publication of 1996 workshop report and conclusion of the 1997 NAAQS 

22 review, the evidence base has continued to expand as described in the 2006 CD and 2013 ISA 

23 (U.S. EPA, 2006; U.S. EPA, 2013). With regard to tree growth effects and crop yield reductions, 

24 results of additional studies conducted in the field have confirmed the tree seedling biomass loss 

25 and crop yield loss concentration-response relationships derived from earlier studies that used 

26 open top chambers (U.S. EPA 2013, section 9.6). Further, similar data, although from an 

 

5 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). 6 For natural ecosystems, they focused on foliar injury as an indicator. For forest ecosystems, 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). 7

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).
6-22 

1 ambient gradient study (outside New York City) rather than a controlled experiment, are 

2 available for an additional, apparently quite sensitive species, the cottonwood (Gregg, 2003). 

3 In the 2008 review, CASAC provided comments related to a cumulative seasonal 

4 secondary standard in the context of their comments on the draft and final Staff Papers and on 

5 the final decision (Henderson, 2006; Henderson, 2007; Henderson, 2008). In all instances, they 

6 conveyed support for establishment of a distinct secondary standard with a cumulative seasonal 

7 form. While the EPA, in the 2007 Staff Paper and 2007 notice of proposed rulemaking, 

8 recognized a broader range of W126 values as appropriate for consideration with regard to a 

9 distinct secondary standard, the CASAC Panel focused on a range they described as 

10 approximately equivalent to that identified by the 1996 workshop participants (Henderson, 2007, 

pp. 3, C-27).8 11 In the CASAC Panel 2006-2007 advice on levels for such a standard, their 

12 suggestion was a focus on levels for a W126 index approximately equivalent to a SUM06 range 

13 of 10 to 20 ppm-hr (Henderson, 2006, 2007, 2008), which they estimated in 2007 to be a range 

14 from 7 (or 7.5) to 15 ppm-hrs. Based on their consideration of the information available in that 

15 review (e.g., with regard to potential magnitude of effects across multiple years), the CASAC 

16 Panel further advised that “[i]f multi-year averaging is employed to increase the stability of the 

17 secondary standard, the level of the standard should be revised downward to assure that the 

18 desired threshold is not exceeded in individual years” (Henderson, 2007, p. 3). The CASAC

19 advice provided on the 2010 proposed reconsideration and thus far in this review is summarized 

20 in section 6.4 below. 

21 In considering the evidence briefly summarized above in the context of levels for a 

22 W126-based standard, we recognize that given the different types of O3-induced effects, genetic 

23 variability within and between species, and environmental modifiers of effects that also 

24 contribute to variability, it is not feasible to identify a range of cumulative seasonal exposures 

25 from the vegetation effects evidence which would provide a consistent degree of protection for 

26 all species. In so doing, we note the importance of considering this evidence in several 

27 dimensions that pertain to judgments required by the Administrator regarding public welfare 

28 significance. For example, we take note of the usefulness of considering the cumulative seasonal 

29 exposure at which the median species response or the majority of the species’ responses are 

30 expected to be below minimal response benchmarks of interest and at which only a very few 

31 species’ responses are expected to exceed more substantial response benchmarks. Before

32 articulating such considerations with regard to specific benchmarks and index values, we first 

33 consider the WREA findings in the context of the following question. 

 

8 Appendix C of the March 26, 2007 CASAC letter 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.
6-23 

1 • What are the nature and magnitude of risks to vegetation estimated for the 

2 average W126 index scenarios evaluated in the WREA, and what is the 

3 magnitude of risk reduction from risks estimated for air quality conditions 

4 estimated for the current standard? 

5 In considering the WREA estimates here we take note of uncertainties in the extent to 

6 which the results for each modeled air quality scenario represent cumulative seasonal O3

7 exposures that would be expected to occur across the three years represented in each scenario. In 

8 general, each scenario is represented by a dataset of 3-year average W126 index values across 

9 the national modeling area. Thus, the results estimated for the various analyses performed do not 

10 reflect any year-to-year variability that would be expected in single year results. Rather, they 

11 reflect average estimates for the three year period modeled. Such estimates do not account for 

12 the potential for compounding of effects in perennial species that has been discussed previously. 

13 Limited analyses in the WREA describe the potential for the WREA estimates to underestimate 

14 cumulative biomass-related effects in perennial species (as noted in sections 6.2 and 5.2.2 above 

15 and described in detail in U.S. EPA, 2014, chapter 6, 6.2.1.4). This potential for underestimation 

16 is recognized in the context of the uncertainties associated with other aspects of the different 

17 analyses in section 6.9 of the WREA (e.g., U.S. EPA, 2014, Table 6-27). We additionally note 

18 that the limited WREA compounding analyses do not take into account other variables which can 

19 affect the magnitude of these effects in the field.

20 In considering the question posed above, we focus particularly on WREA estimates 

21 related to O3 effects on plant biomass and associated ecosystem services effects. In so doing, we 

22 note the relationships among effects on individual plants to other ecosystem components and 

23 functions, such as carbon sequestration and air pollutant removal (U.S. EPA 2013, section 

24 9.4.3.4; U.S. EPA, 2014, sections 6.6 and 6.7), as well as market responses to changes in timber 

25 and agricultural production (U.S. EPA, 2014, sections 6.3 and 6.5). We additionally recognize 

26 other biomass-related responses, such as non-timber forest products, and other O3-attributable 

27 ecosystem responses for which we have primarily qualitative characterizations of impacts (U.S. 

28 EPA, 2014, chapter 5).

29 We turn first to the WREA estimates for a range of effects related to biomass loss, which 

30 are based on application of the C-R functions for seedlings of 12 tree species described in the 

31 ISA (U.S. EPA, 2013, section 9.6.2) and the WREA (U.S. EPA, 2014, section 6.2). First we 

32 note (as considered above) the range of responses for the individual species for which C-R 

33 functions have been developed. These twelve species vary appreciably in sensitivity of growth 

34 reduction (in terms of relative biomass loss, or RBL) in response to O3 exposure. For example, 

35 2% seedling biomass loss is estimated to occur following exposure represented by W126 index 

36 values below 10 ppm-hours for seven of the studied species and at or above approximately 20 
6-24 

1 ppm-hours in the other five studied species (Figure 5-1 above). The WREA presents three 

2 approaches for characterizing a median response across the species (Figure 5-2; U.S. EPA 2014, 

3 section 6.2.1.2). Across the three approaches, the median seasonal W126 index value for which 

4 a 2% biomass loss is estimated in seedlings for the studied species ranges between approximately 

5 7 and 14 ppm-hrs (see description in section 5.2.1 above). 

6 We additionally consider the WREA estimates of overall ecosystem-level effects from 

7 biomass loss considering the 12 studied species together (U.S. EPA 2014, section 6.8). The 

8 WREA analysis used the species-specific biomass loss C-R functions, information on prevalence 

9 of the studied species across the U.S., and a weighting approach based on proportion of the basal 

10 area within each grid cell that each species contributes. A weighted RBL value for each grid cell 

11 is generated by weighting the RBL value for each studied tree species found within that grid cell 

12 by the proportion of basal area it contributes to the total basal area of tree species within the grid 

13 cell, and then summing those individual weighted RBLs. Based on the average W126 values 

14 estimated for the air quality scenario just meeting the current standard across the contiguous 

15 U.S., the WREA estimates 0.8 percent of the total geographic area to have a weighted relative 

16 biomass loss above 2% (Table 6-2 below; U.S. EPA 2014, Table 6-24). In the W126 air quality 

17 scenarios for 15, 11, and 7 ppm-hrs (average across three years), the percent of total area having 

18 weighted relative biomass loss greater than two percent was 0.7 percent, 0.5 percent and 0.2 

19 percent, respectively (Table 6-2 below; U.S. EPA 2014, Table 6-25). In considering these 

20 estimates, we note that the values for percentages of basal area include many grid cells in which 

21 none of the 12 studied species are found. Thus, since only 64 percent of grid cells contained one 

22 or more of the 12 species, these values are likely to be low and illustrate the potential 

23 uncertainties associated with the limited number studied tree species for which C-R functions 

24 exist (U.S. EPA, 2014, section 6.8). 

25 
6-25 

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

2 loss in WREA air quality scenarios. 

----------------Air Quality Scenarios ---------------------------

Conditions just 

meeting the current 

standardA

W126 index scenariosB

15 ppm-hrs 11 ppm-hrs 7 ppm-hrs

Percent of total area with 

wRBL>2% 0.8 % 0.7 % 0.5 % 0.2 %

A This analysis uses air quality values that are estimated per model grid cell using the W126 value assigned to the grid cell 

based on application of the VNA method to the monitor-location W126 values that are the average at that location across 

the 3 years of W126 values for the model-adjusted dataset that just meets the current standard (fourth-highest daily 

maximum 8-hour concentration, averaged over 3 consecutive years of 75 ppb).

B The national distribution of W126 values within model grid-cells for each scenario reflects model 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).

3 

4 We next consider the wRBL estimates from the WREA analysis of 145 (of the 155) 

5 federally-designated Class I areas for which there was sufficient information regarding O3-

6 sensitive species (U.S. EPA, 2014, section 6.8.1, Table 6-26, Appendix 6E). These 145 parks 

7 had at least one O3-sensitive tree species for which a C-R function for RBL was available. Using 

8 the C-R functions for the species found within each park, the WREA calculated an average 

9 wRBL value for each park for the 3-year average W126 index values estimated in those locations 

10 for the current standard and three W126 air quality scenarios. Under conditions model-adjusted 

11 to just meet the current standard, the average wRBL in 2 of the 145 parks is estimated to be 

12 above 2%, as presented in Table 6-3 below. We compare this estimate to those for the W126 

13 scenarios. For the W126 scenarios of 15 and 11 ppm-hrs, the estimated weighted RBL is greater 

14 than 2% in two of the 145 parks, while it is great than 2% in only 1 park for the 7 ppm-hrs 

15 scenario. 

16 

17 
6-26 

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

2 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 value assigned to the 

grid cell based on application of the VNA method to the monitor-location W126 values that are the average at that location across 

the 3 years of W126 values for the model-adjusted dataset that just meets the current standard (fourth-highest daily maximum 8-

hour concentration, averaged over 3 consecutive years of 75 ppb).

B The national distribution of W126 values within model grid-cells for each scenario reflects model 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).

3 

4 With respect to crops, based on the C-R functions described in the ISA and additionally 

5 analyzed in the WREA, a 5% median relative yield loss (RYL) for all studied crop cases occurs 

6 within the range of W126 index values from 12-17 ppm-hrs (U.S. EPA 2014, Figure 6-5). The 

7 species-specific C-R functions project less than 5% yield loss for 6 out of 10 species for O3

8 exposure equivalent to a seasonal W126 index value of 15 ppm-hrs and for all of the 10 studied 

9 species after an exposure equivalent to a seasonal W126 index of 7 ppm-hrs (Table 6-1 above, 

10 U.S. EPA 2014, section 6.2 and Appendix 6A). For soybeans, less than 5% yield loss was 

11 estimated for the W126 index value of 12 ppm-hrs (U.S. EPA 2014, Figure 6-3). The WREA 

12 estimates of crop yield loss for the modeled air quality scenarios are summarized in Table 6-4 

13 below (details are provided in U.S. EPA 2014, section 6.5.1 and Appendix 6B). For the recent 

14 air quality conditions scenario, the means for all crops were less than 5% loss across all states. 

15 Crop yield loss estimates for all states were also less than 5% in the air quality scenario 

16 representing conditions just meeting the current standard (U.S. EPA, 2014, section 6.5.1 and 

17 Appendix 6B). 

18 
6-27 

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

2 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 value assigned to the cell based on 

application of the VNA method to the monitor-location W126 values that are the average at that location across the 3 years of 

W126 values for the model-adjusted dataset that just meets the current standard (fourth-highest daily maximum 8-hour 

concentration, averaged over 3 consecutive years of 75 ppb).

B The national distribution of W126 values within grid cells for each scenario reflects model 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).

3 

4 As explained above, however, comparisons of the WREA’s air quality scenarios for the 

5 national-scale estimates of timber production and consumer and producer surpluses are not 

6 straightforward to interpret due to market dynamics. Estimates for the recent conditions and 

7 current standard scenarios are compared to the three W126 scenarios. In general, substantially 

8 greater economic surpluses (approximately 51 billion in terms of 2010 dollars) are estimated 

9 from the comparison of the recent conditions (2006-2008) scenario to the current standard 

10 scenario. The vast majority of these economic surpluses are estimated for agricultural 

11 production. Differences of the average W126 scenarios from the current standard scenario are 

12 much smaller (U.S. EPA 2014, Appendix 6B). 

13 Because increases in timber production represent increased tree growth and concurrent 

14 carbon sequestration, we also consider WREA estimates of the potential increase in carbon 

15 storage that potentially could occur for different air quality scenarios (U.S. EPA 2014, section 

16 6.6.1). Comparisons of the W126 scenarios to the current standard scenario with regard to 

17 carbon sequestration estimates do not indicate an appreciable difference for the W126 scenario 

18 of 15 ppm-hrs. The majority of the enhanced carbon sequestration potential in the forest biomass 

19 increases over time, for alternatives, is predicted to occur for the W126 scenarios of 11 and 7 

20 ppm-hrs. Over 30 years, the current standard scenario projection is 89,184 million metric tons of 

21 CO2 equivalents (MMtCO2e). The WREA estimates additional sequestration potential of 13, 593 
6-28 

1 and 1,600 MMtCO2e, for the W126 scenarios of 15, 11 and 7 ppm-hrs, respectively, as compared 

to the current standard baseline (U.S. EPA 2014, Table 6-18)9 2 . We also take note of the 

3 relatively smaller estimates for carbon sequestration associated with improved crop yields (over 

4 30 years) in the agricultural sector, which indicate little difference among the different W126 

5 scenarios. 

6 We additionally consider the WREA estimates for five urban areas of how reduced 

7 growth of O3-sensitive trees in urban forests may affect the ecosystem services of air pollutant

8 removal and carbon sequestration (U.S. EPA, 2014, sections 6.6.2 and 6.7 and Appendix 6D). 

9 With regard to air pollutant removal, the WREA estimated metric tons of carbon monoxide, 

10 nitrogen dioxide, ozone and sulfur dioxide removed under the modeled air quality scenarios. In 

11 considering these estimates we note the general assumptions made to estimate order of 

12 magnitude effects of O3 removal by trees on O3 concentrations in the five urban areas and the 

13 associated uncertainties (U.S. EPA 2014, sections 6.7 and 6.9 and Appendix 6D). Estimates for 

14 all five areas indicate increased pollutant removal for air quality model-adjusted from recent 

15 conditions to just meet the current standard, with much smaller difference where they exist 

16 between the current standard and three W126 scenarios (Table 6-5 below). With respect to 

17 carbon sequestration, relative to the scenario representing just meeting the current standard, 

18 again the largest difference is generally observed with the existing conditions scenario (Table 6-

19 5). This is because of the largest difference in W126 estimates occurring between these two 

20 scenarios. In addition to the small differences in W126 index values among the current standard 

21 and W126 scenarios for these five areas, which contribute to the similarities among scenarios, we 

22 also note that as only 2 or 3 tree species were able to be assessed in each city, these results may 

23 underestimate the overall impacts nationally, although other areas of uncertainty (recognized 

24 below) may tend to contribute to the opposite potential (U.S. EPA 2014, Table 6-27).

25 

 

9 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 2013and available at http://www.epa.gov/cleanenergy/energyresources/calculator.html). 
6-29 

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

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

3 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 15 ppm-hrs 15 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

Million Metric Tons of CO2 Equivalents, Carbon Storage (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 value 

assigned to the grid cell based on application of the VNA method to the monitor-location W126 values that are the average at 

that location across the 3 years of W126 values for the model-adjusted dataset that just meets the current standard (fourthhighest daily maximum 8-hour concentration, averaged over 3 consecutive years of 75 ppb).

B The national distribution of W126 values within model grid-cells for each scenario reflects model 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).

4 

5 With regard to foliar injury, we take note of the WREA analyses of the nationwide 

6 dataset (2006- 2010) for U.S. Forest Service biosites described in section 5.4.2 above, including 

7 the presentation indicating that the proportion of biosites with injury varies with soil moisture 

8 conditions and O3 W126 index values, and that the proportion of biosites with injury severity 

9 greater than 5 also varied with soil moisture (U.S. EPA 2014, Chapter 7, Figure 7-11). The 

10 evidence of O3-attributable visible foliar injury incidence occurring in USFS biosites shows that 

11 the proportion of biosites showing foliar injury incidence increases steeply with W126 index 

12 values up to approximately10 ppm-hrs. At W126 index levels greater than approximately 10 

13 ppm-hrs the proportion of sites showing foliar injury incidence is relatively constant.

14 In reflecting across the range of W126 estimates identified in various WREA analyses, 

15 we first note the substantial benefits and reductions in biomass-related risks estimated for air 
6-30 

1 quality adjusted to just meet the current standard. Additional incremental benefits or risk 

2 reduction, generally of relatively smaller magnitude, is estimated across the W126 scenarios. In 

3 considering this information discussed above in the context of identifying levels appropriate to 

4 consider for a W126-based standard, we first take note of associated uncertainties in the context 

5 of the following question. 

6 • What are important uncertainties and limitations in the evidence and 

7 exposure/risk analyses?

8 In considering the evidence and exposure/risk information summarized above and the 

9 weight to place on this information, we are mindful of the uncertainties and limitations 

10 associated with several key aspects of this information. We first consider the uncertainties 

11 associated with the evidence underlying the tree seedling and crop C-R functions, given the 

12 importance of these functions for many of the ecosystem service analyses described in the 

13 WREA. Several key uncertainties associated with this information are listed here.

14 • Uncertainty regarding the extent to which the subset of studied tree and crop species 

15 encompass the total number of O3 sensitive species in the nation and to what extent it 

16 is representative of U.S. vegetation as a whole, given that information is available for 

17 only a small fraction of the number of total species of trees and crops grown in the 

18 U.S. (U.S. EPA, 2013, section 9.6, U.S. EPA, 2014, Table 6-27). 

19 • Uncertainties regarding intra-species variability due to the different numbers of studies 

20 that exist for different species so that the weight of evidence is not the same for each 

21 species. Those species with more than one study show variability in response and C-R 

22 functions. The potential variability in less well studied species is therefore unknown 

23 (U.S. EPA, 2013, pp. 9-123/125, U.S. EPA, 2014, section 6.2.1.2, and Table 6-27). 

24 • Uncertainty regarding the extent to which tree seedling C-R functions can be used to 

25 represent mature trees since seedling sensitivity has been shown in some cases to not 

26 reflect mature tree O3 sensitivity in the same species (U.S. EPA, 2013, section 9.6, 

27 U.S. EPA, 2014, section 6.2.1.1 and Tables 6-5 and 6-27). 

28 • Uncertainty in the relationship of O3 effects on tree seedlings (e.g., relative biomass 

29 loss) in one or a few growing seasons to effects that might be expected to accrue over 

30 the life of the trees extending into adulthood (U.S. EPA, 2013, pp. 9-52/53, U.S. 

31 EPA, 2014, section 6.2.1.4 and Table 6-27). 

32 • Uncertainties associated with estimating the national scale ecosystem-level impacts 

33 using weighted relative biomass loss (U.S. EPA, 2014, section 6.8, and Table 6-27) 

34 • Uncertainties associated with potential biomass loss in federally designated Class I 

35 areas (U.S. EPA, 2014, section 6.8. and Table 6-27) 

36 Turning to consideration of the air quality conditions estimated for the various air quality 

37 scenarios, we take note of the following uncertainties associated particularly with estimates of O3

38 exposures in rural areas nationally. These are described more completely in chapter 4 of the 
6-31 

1 WREA (see for example, U.S. EPA, 2014, section 4.4) and summarized in chapter 8 of the REA 

2 (U.S. EPA, 2014, section 8.5). 

3 • Uncertainties in O3 exposures due to a lack of rural monitors, especially in the western 

4 U.S. and at high elevation sites.

5 • Uncertainties associated with the method (VNA) used to interpolate monitor values to 

6 estimate W126 index values in locations without monitors. 

7 • Uncertainties in model-adjusted estimates of O3 concentrations associated with 

8 meeting the current standard and potential alternative W126-based standards.

9 Numerous ecosystem services assessments were described in the WREA. These 

10 assessments relied heavily on models, which also relied on the inputs of the seedling and crop C11 R functions and model-adjusted air quality estimates. Thus, including the uncertainties from the 

12 first two categories discussed above, additional uncertainties associated with the ecosystem 

13 services models include the following. 

14 • Uncertainties associated with use of the iTree model to estimate pollution removal and 

15 carbon storage in 5 urban area case studies, including uncertainties in the base 

16 inventory of city trees, the functions used for air pollution removal and carbon storage 

17 (U.S. EPA, 2014, sections 6.6.2, 6.7, and Table 6-27).

18 • Uncertainties associated with use of the FASOMGHG model for national timber and 

19 crop production, including use of median C-R functions for crops in FASOM and 

20 crop proxy and forest type assumptions to fill in where there was insufficient data 

21 (U.S. EPA, 2014, sections 6.3, 6.5, 6.6.1, and Table 6-27).

22 • Uncertainties associated with use of the FASOMGHG model to estimate national scale 

23 carbon sequestration, including those associated with the functions for carbon 

24 sequestration (U.S. EPA, 2014, sections 6.2.1.1, 6.6.1, and Table 6-27). 

25 In addition to biomass loss and crop yield loss, the WREA estimates the incidence and/or 

26 severity of O3-induced visible foliar injury, both at the national and national park scales. 

27 Numerous uncertainties are associated with these assessments and include the following. 

28 • Uncertainties associated with our understanding of the number and sensitivity of O3

29 sensitive species (U.S. EPA, 2014, sections 7.2.1, 7.5 and Table 7-22). 

30 • Uncertainties associated with spatial assignment of foliar injury biosite data to 12x12 

31 km grids (U.S. EPA, 2014, sections 7.2.1, 7.5 and Table 7-22). 

32 • Uncertainties associated with availability of biosite sampling data in some locations in 

33 the western U.S. (U.S. EPA, 2014, sections 7.2.1, 7.5 and Table 7-22). 

34 • Uncertainties associated with soil moisture threshold for foliar injury (U.S. EPA, 

35 2014, sections 7.2.2, 7.2.3, 7.5 and Table 7-22). 
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1 • Uncertainties associated with spatial resolution of soil moisture data, time period for 

2 soil moisture data, drought categories and the combination of soil moisture and 

3 biosite data (U.S. EPA, 2014, sections 7.3.3.2, 7.5 and Table 7-22). 

4 • Uncertainties associated with O3 exposure data of vegetation and recreational areas 

5 within parks (U.S. EPA, 2014, sections 7.4, 7.5 and Table 7-22). 

6 • Uncertainties associated with surveys of recreational activities (U.S. EPA, 2014, 

7 sections 7.1.1.2, 7.5 and Table 7-22). 

8 Additionally, there is uncertainty associated with the extent to which the endpoints and 

9 associated risk estimates considered above represent effects reasonably judged adverse in the 

10 context of public welfare. All of these uncertainties are important to considerations below in the 

11 context of target levels of protection with regard to weight to be placed on various lines of 

12 evidence and assessment results.

13 • What considerations may be important to the Administrator’s judgments on the 

14 public welfare significance of O3 associated vegetation effects that may be 

15 expected under air quality conditions associated with different levels for a 

16 seasonal cumulative standard? 

17 Our consideration of this question is intended to provide a public welfare context for 

18 consideration of the evidence and exposure/risk information discussed above, which includes the 

19 nature and magnitude of observed and predicted effects at various levels of cumulative seasonal 

20 exposures. We also note the importance of considering information in an integrated manner, 

21 rather than focusing only on results from any one analysis. For example, we find it appropriate, 

22 in considering the evidence with regard to seedling growth reduction (or biomass loss), to 

23 consider the WREA estimates of affected area based on tree basal area together with estimates of 

24 individual species responses based simply on the evidence-based C-R functions, and in light of 

25 other potential impacts summarized above. In so doing in section 6.5 below, we take into 

26 account considerations relevant to public welfare policy judgments required of the 

27 Administrator, such as those described here. 

28 As recognized in sections 1.3.2 and 5.1, the Clean Air Act specifies that secondary 

29 standards protect against known or anticipated adverse effects to public welfare. In the 

30 Administrator’s judgment as to the standards that would be requisite (i.e., neither more nor less 

31 stringent than necessary) to protect the public welfare under the Act, she will consider a number 

32 of factors including 1) what can be considered to constitute an adverse effect to the public 

33 welfare; 2) the nature and magnitude of the effects and the risks that remain after meeting the 

34 level of the current standard; and, 3) what is necessary to achieve the requisite (no more and no 

35 less) degree of protection. In the 2008 decision by which the current standard was established, 

36 the Administrator considered these factors in judging the previously-existing standard to not 

37 provide the requisite public welfare protection. At that time the Administrator found that the 
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1 exposure- and risk-based analyses available in that review indicated that adverse effects to 

2 vegetation would be predicted to occur under air quality conditions associated with just meeting 

3 the then-current standard. The effects identified were “visible foliar injury and seedling and 

4 mature tree biomass loss in O3-sensitive vegetation” (73 FR 16496). In so noting, the 

5 Administrator indicated that he believed that “the degree to which such effects should be 

6 considered to be adverse depends on the intended use of the vegetation and its significance to 

7 public welfare” (73 FR 16496). With regard to consideration of intended use, the Administrator 

8 took note of the specific uses of public lands set aside by Congress and intended to provide 

9 benefits to the public welfare, “including lands that are to be protected so as to conserve the 

10 scenic value and the natural vegetation and wildlife within such areas, and to leave them 

11 unimpaired for the enjoyment for future generations” such as Class I areas (73 FR 16496). The 

12 Administrator also recognized areas set aside by States, Tribes and public interest groups with 

13 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).10 14 

15 In the Administrator’s judgments in the 2008 review, he did not identify specific criteria 

16 or benchmarks or an overall level of protection from adverse environmental effects to public 

welfare judged to be requisite under the Act.11 17 As noted above, the scientists at the 1996 

18 workshop identified ranges of cumulative seasonal index values (e.g., in terms of SUM06 or 

19 W126) in the context of considering a degree of protection for vegetation effects defined in terms 

20 of relative yield loss in crops and relative biomass loss in tree seedlings. In considering this 

21 information in the context of a secondary standard, judgments are required by the Administrator 

22 with regard to the degree that these or other benchmarks and other effects should be judged 

23 adverse to the public welfare. In considering levels for a W126-based secondary standard that 

24 may be appropriate to consider, we recognize that the statute requires that a secondary standard 

25 be protective against only those known or anticipated O3 effects that are “adverse” to the public 

26 welfare, not all identifiable O3-induced effects. Thus, we recognize both the importance of 

27 scientific consensus statements that have been made regarding vegetation-related endpoints and 

28 O3 exposure levels that might protect against such key endpoints and the importance of placing 

29 such conclusions in the context of consideration of the public welfare more broadly. 

 

10 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). 11 In remanding the 2008 decision on the secondary standard back to EPA, the Court indicated this 

omission, as described in section 1.2.2 above.
6-34 

1 As discussed in section 5.1 and recognized by the EPA in prior reviews, staff recognizes 

2 the importance of a more expansive construct or paradigm that addresses what constitutes 

3 adverse effects of O3 to public welfare. In so doing, we also recognize several aspects or 

4 dimensions of vegetation effects for consideration within this paradigm. These include the 

5 likelihood, type, magnitude, and spatial scale of the effect, as well as the potential for recovery 

6 and any uncertainties relating to these conditions (77 FR 20231). As in the last review, we also 

7 continue to recognize that the public welfare significance of O3-induced effects on sensitive 

8 vegetation growing within the U.S. can vary, depending on the nature of the effect, the intended 

9 use of the sensitive plants or ecosystems, and the types of environments in which the sensitive 

10 vegetation and ecosystems are located. Any given O3-related effect on vegetation and 

11 ecosystems (e.g., biomass loss, foliar injury), therefore, may be judged to have a different degree 

12 of impact on the public welfare depending, for example, on whether that effect occurs in a Class 

13 I area, a city park, or in commercial cropland. In the 2008 review, the Administrator judged it 

14 appropriate that this variation in the significance of O3-related vegetation effects should be taken 

15 into consideration in judging the level of ambient O3 that is requisite to protect the public welfare 

16 from any known or anticipated adverse effects (73 FR 16496). For example, in considering 

17 visible foliar injury and seedling and mature tree biomass loss in O3-sensitive vegetation

18 expected under alternative air quality scenarios, the Administrator noted that “the degree to 

19 which such effects should be considered to be adverse depends on the intended use of the 

20 vegetation and its significance to the public welfare” (73 FR 16496). Further, the rulemaking 

21 notice stated that “[i]n considering what constitutes a vegetation effect that is adverse from a 

22 public welfare perspective, the Administrator believes it is appropriate to continue to rely on the 

23 definition of ‘adverse,’ ... that imbeds the concept of ‘‘intended use’’ of the ecological receptors 

24 and resources that are affected, and applies that concept beyond the species level to the 

25 ecosystem level.” The notice went on to state that “[i]n so doing, the Administrator has taken 

26 note of a number of actions taken by Congress to establish public lands that are set aside for 

27 specific uses that are intended to provide benefits to the public welfare, including lands that are 

28 to be protected so as to conserve the scenic value and the natural vegetation and wildlife within 

29 such areas, and to leave them unimpaired for the enjoyment of future generations” (73 FR 

30 16496). Such public lands that are protected areas of national interest include national parks and 

31 forests, wildlife refuges, and wilderness areas.

32 We also consider effects on ecosystem services in considering adversity to public 

33 welfare. For example, the WREA has evaluated the economic value of ecosystem services 

34 affected by O3 and how those services might be expected to change under different air quality 

35 scenarios representing the current and potential alternative standards (U.S. EPA, 2014, chapter 

6). 36 
6-35 

1 Thus, we recognize several important considerations in evaluating levels of protection 

2 and levels for a cumulative seasonal W126-based standard including: the extent of areas 

3 expected to be affected nationwide and the magnitude of those effects; the extent of effects in 

4 areas of national significance; the extent to which these impacts might be judged significant from 

5 a public welfare perspective and associated uncertainties in the information. Accordingly, we 

6 recognize that the range of alternative standard levels that may be appropriate to consider differs 

7 based on the weight placed on different aspects of the evidence and on different aspects of the 

8 quantitative exposure/risk information, and the associated uncertainties, as well as on public 

9 welfare policy decisions regarding the public welfare significance of the effects considered and 

10 the approaches for considering benchmarks for growth or biomass loss and other vegetation 

11 effects of O3. As described in chapter 1, our objective is to identify the range of policy options 

12 supported by the current evidence- and exposure/risk-based information and with consideration 

13 of the role of the Administrator’s public welfare judgments. In so doing, we recognize support 

14 for consideration of a broad range of W126 index values, which we discuss in section 6.5, with 

15 recognition of the different judgments that might provide support for different parts of such a 

16 range.

17 6.4 CASAC ADVICE

18 In our consideration of potential alternative standards, in addition to the evidence-based,

19 risk/exposure-based, and air quality information discussed above, we also consider the advice 

20 and recommendations of CASAC in EPA’s proposed 2010 reconsideration of the 2008 decision, 

21 as well as comments received thus far in the current review, in the context of its review of the 

22 ISA, and earlier drafts of this document and the WREA. We have additionally considered public 

23 comments received to date, some of which have suggested a lack of new information for support 

24 of a distinct secondary standard and others that urge the consideration of a secondary standard 

25 with a cumulative seasonal form using the W126 metric and a level within the range of 7 to 15 

ppm-hrs.12 26 

27 In response to the EPA’s solicitation of CASAC’s advice on the Agency’s proposed 

rulemaking as part of the reconsideration,13 28 CASAC conveyed their support for a secondary 

29 standard distinct from the primary standard, noting that “vegetation effects are more dependent 

 

12 Public comment received thus far in this review are in the docket EPA–HQ–OAR-2008-0699, accessible 

at www.regulations.gov. 13 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-hour (75 FR 3027). 
6-36 

1 on the cumulative exposure to, and uptake of, ozone over the course of the entire growing 

2 season” (Samet, 2010). 

3 CASAC also supports EPA’s secondary ozone standard as proposed: a new 

4 cumulative, seasonal standard expressed as an annual index of the sum of 

5 weighted hourly concentrations (i.e., the W126 form), cumulated over 12 hours 

6 per day (8am to 8pm) during the consecutive 3-month period within the ozone 

7 season with the maximum index value, set as a level within the range of 7 to [1]5 

8 ppm-hours. This W126 metric can be supported as an appropriate option for 

9 relating ozone exposure to vegetation responses, such as visible foliar injury and 

10 reductions in plant growth. We found the Agency’s reasoning ... to be supported 

11 by the extensive scientific evidence considered in the last review cycle. 

12 In advice offered so far in the current review, which considers an updated scientific and 

13 technical record since the 2008 rulemaking, the CASAC stated that “the focus [of the first draft 

14 PA] on the W126 form is appropriate” (Frey and Samet, 2012, p. 2). They further commented 

15 with regard to the support provided in the first draft PA for the consideration of such a form for a 

16 secondary standard (Frey and Samet, 2012, p. 2). 

17 There is a strong justification made for using a cumulative and weighted exposure 

18 standard for welfare effects (i.e. the W126), and for the utility of using a 3-month 

19 daylight exposure metric. Averaging across years is not recommended because a 

20 single high exposure year could have lasting effects because of the perennial 

21 nature of many plants and the lag times associated with propagating effects 

22 through ecosystem trophic levels. Averaging would obscure such critical impacts 

23 and lead to inadequate protection against welfare effects. 

24 Additionally, in advice regarding consideration of levels, their advice indicated that “[o]ptions 

25 for levels [of a secondary standard] should be based on factors including predicted 5% loss of 

26 crop yield and predicted 1-2% loss for trees” (Frey and Samet, 2012, p. 2). 

27 6.5 PRELIMINARY STAFF CONCLUSIONS ON ALTERNATIVE STANDARD

28 Staff’s consideration of alternative secondary O3 standards builds on our conclusion from 

29 section 5.7 above that the body of evidence, in combination with the results of the WREA, calls 

30 into question the adequacy of the current secondary standard and provides support for 

31 consideration of alternative standards. In sections 6.1 to 6.3 above, we consider how the 

32 currently available scientific evidence and exposure/risk information inform decisions regarding 

33 the basic elements of the NAAQS: indicator (6.1), form and averaging time (6.2), and level (6.3). 

34 In so doing, we consider both the information available at the time of the last review and 

35 information newly available since the last review which has been critically analyzed and 

36 characterized in the 2013 ISA. As an initial matter, with regard to the indicator, we conclude 

37 that based on the available science it is still appropriate to continue to use measurements of O3 in 
6-37 

1 accordance with federal reference methods as the indicator to address effects associated with 

2 exposure to ambient O3 alone or in combination with related photochemical oxidants. 

3 In considering alternative standards, staff has considered the available body of evidence 

4 as comprehensively assessed in the ISA, the risk and exposure information presented in the 

5 second draft REA and CASAC advice and public comment thus far in this review with regard to 

6 support for consideration of options that are different from the current standard, as articulated by 

7 the following overarching question 

8 • To what extent does the currently available scientific evidence- and exposure/risk 

9 based information, as reflected in the ISA and WREA, support consideration of 

10 alternatives to the current O3 standard to provide increased protection from 

11 ambient O3 exposures?

12 In considering potential forms alternative to that of the current standard, we note that the 

form for the current secondary standard is the 4th 13 highest daily maximum 8-hour average, 

14 averaged over three years. As discussed in chapter 5 and section 6.2 above, the longstanding 

15 evidence regarding the fundamental aspects of O3 exposure that are directly responsible for 

16 inducing vegetation response indicates that plant response to O3 is driven by the cumulative 

17 exposure to O3 during the growing season, rather than by a single event (U.S. EPA, 2013, section 

18 2.6.6.1). This cumulative exposure depends on both the total duration of the exposure (from 

19 repeated O3 episodes) and the concentrations of those exposures (higher concentrations having a 

20 disproportionate impact over lower concentrations). On the basis of this longstanding and 

21 extensive evidence, the ISA concludes that exposure indices that cumulate and differentially 

22 weight the higher hourly average concentrations over a season and also include the mid-level 

23 values offer the most defensible approach for use in developing response functions and in 

24 defining indices for vegetation protection (U.S. EPA, 2013, section 2.6.6.1). 

25 CASAC advice in the 2008 review and on the 2010 proposed reconsideration has 

26 additionally recognized that the nature of the exposures relevant to vegetation response is well 

27 described by a cumulative seasonal form and has supported the use of such a form for a 

28 secondary O3 standard (Henderson, 2006; Samet, 2010). The current CASAC O3 Panel has 

29 expressed similar views, stating “[t]here is a strong justification made for using a cumulative and 

30 weighted exposure standard for welfare effects (i.e. the W126)...” (Frey and Samet, 2012, p. 2). 

31 We also note that on the basis of the evidence and exposure/risk information available in the two 

32 previous reviews, and in consideration of CASAC advice, the Administrator has recognized the 

33 importance of protecting vegetation from cumulative, seasonal exposures and proposed such a 

34 form as an appropriate reasonable policy option (61 FR 65741-44; 62 FR 37899-905; 75 FR 

35 3012-3027). 
6-38 

1 Thus, in considering alternative forms of the standard we conclude that it is reasonable 

2 and appropriate to consider a cumulative, concentration-weighted form to provide protection 

3 against cumulative, seasonal exposures to O3 that are known or anticipated to harm sensitive 

4 vegetation or ecosystems. Such a form is specifically designed to directly measure the kind of 

5 O3 exposures that can cause harm to vegetation and would have a distinct advantage over the 

6 form of the current standard in characterizing air quality conditions potentially of concern for 

7 vegetation and demonstrating that the desired degree of protection against those conditions was 

8 being achieved. 

9 In considering the appropriate index for a cumulative seasonal form, we recognize that a 

10 number of different cumulative concentration weighted indices have been developed and have 

11 been evaluated in the scientific literature and in past NAAQS reviews in terms of their ability to 

12 predict vegetation response and their usefulness in the NAAQS context (U.S. EPA, 2006, pp. 9-

13 11 to 9-15 and pp. AX9-159 to AX9-187; U.S. EPA, 2007, pp. 7-15/16). While these various 

14 forms have different strengths and limitations, as noted in the ISA (U.S. EPA, 2013, section 9.5), 

the W126 index14 15 has some important aspects not shared by other non-sigmoidally weighted 

16 cumulative indices. For example, given the lack of discernible threshold for vegetation effects in 

17 general, we recognize the fact that the W126 metric does not have a cut-off in its weighting 

18 scheme (down to about 30 ppb below which the weighting factor is effectively zero), such that it 

19 includes consideration of potentially damaging lower O3 concentrations. Additionally, the W126 

20 metric also adds increasing weight to hourly concentrations from about 40 ppb to about 100 ppb, 

21 an important feature because “as hourly concentrations become higher, they become increasingly 

22 likely to overwhelm plant defenses and are known to be more detrimental to vegetation” (U.S. 

23 EPA, 2013, p. 9-104). We additionally take note of CASAC advice in the 2008 review and on 

24 the 2010 proposed reconsideration recommending the use of the W126 index for a cumulative 

25 seasonal form for a secondary O3 standard (Henderson, 2006; Samet, 2010). Similarly, the 

26 current CASAC O3 Panel has indicated that a focus on a W126 form is appropriate (Frey and 

27 Samet, 2012) Therefore, on the basis of the strength of the evidence and advice from CASAC, 

28 we conclude that the W126 index is the most appropriate cumulative seasonal form to consider 

29 in the context of the secondary O3 NAAQS review. 

30 We next turn to the exposure periods – diurnal and seasonal – over which the W126 

31 index would be summed in any given year. As discussed in section 6.2 above, the currently 

32 available information continues to provide support for a definition of the diurnal period of 

 

14 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).
6-39 

1 interest as the 12-hour period from 8 am to 8 pm (U.S. EPA, 2013, section 9.5.3). In prior 

2 reviews, the EPA has identified the 12-hour period from 8 am to 8 pm as appropriately capturing 

3 the diurnal window with most relevance to the photosynthetic process (72 FR 37900; 75 FR 

4 3013) and CASAC has generally supported the 12 hour daylight period (Henderson, 2006, 2007). 

5 In light of the continued support in the evidence base, and no evidence on this issue differing 

6 from that in previous reviews, we again conclude that it is appropriate to use the 12-hour period 

7 from 8 am to 8 pm to cumulate daily O3 exposures. On this basis, we conclude that the 12-hour 

8 diurnal window (8:00 am to 8:00 pm) represents the portion of the diurnal exposure period that is 

9 most relevant to predicting or inducing plant effects related to photosynthesis and growth and 

10 thus is an appropriate diurnal period to use in conjunction with a W126 cumulative metric. 

11 With regard to a seasonal period of interest, the current evidence base continues to 

12 provide support for a seasonal period with a minimum duration of three months (U.S. EPA, 

13 2013, section 9.5.3). Included in the currently available evidence is a new analysis that 

14 compared 3- and 6-month maximum W126 values for over 1,200 AQS and CASTNET EPA 

15 monitoring sites for the years 2008-2009 that found that the two accumulation periods were 

16 highly correlated (U.S. EPA, 2013, section 9.5.3, Figure 9-13). Thus, although we recognize that 

17 the selection of a single seasonal time period over which to cumulate O3 exposures for a national 

18 standard necessarily represents a balance of factors, given the significant variability in growth 

19 patterns and lengths of growing season among vegetative species growing within the U.S., we 

20 conclude it is appropriate to identify the seasonal W126 index value as that derived from the 

21 consecutive 3-month period within the O3 season with the highest W126 index value. We note 

22 that such a 3-month exposure period was also supported by CASAC in advice provided during 

23 the last review and on the 2010 proposed reconsideration (Henderson, 2006; Samet, 2010). 

24 With regard to form, we additionally consider the period of time over which a cumulative 

25 seasonal W126-based standard should be evaluated. In so doing, we have considered the support 

26 for both a single year standard and one with a form averaged over three years (section 6.2). We 

27 recognize that there are a number of O3-induced effects that have the potential for public welfare 

28 significance within the annual timeframe (i.e. reduced crop yields and visible foliar injury). 

29 However, as noted in section 6.2 above, there are uncertainties associated with these effects that 

30 make it difficult to determine the degree of annual protection needed to protect the public 

31 welfare from any known or anticipated adverse effects. On the other hand, annual effects in 

32 perennial species can be “carried over” into the subsequent year where they affect growth and 

33 reproduction (U.S. EPA, 2013, pp. 9-43 to 9-44 and p. 9-86). When these annual effects occur 

34 over multiple years due to elevated O3 exposures across several years, they have the further 

35 potential to be compounded, increasing the potential for effects at the population and ecosystem 

36 level, including effects on associated ecosystem services that may be of greater significance to 
6-40 

1 the public welfare. These impacted services can include alteration of below-ground 

2 biogeochemical cycles, and alteration of both above- and below- ground terrestrial community 

3 composition and terrestrial ecosystem water cycling (U.S. EPA, 2013, Table 9-19) and 

4 reductions in productivity and carbon sequestration in terrestrial ecosystems. We additionally 

5 note that multiple years of critical O3 exposures might be expected to result in larger impacts on 

6 forested areas, i.e. increased susceptibility to other stressors such as insect pests, disease, co7 occurring pollutants and harsh weather, than intermittent occurrences of such exposures due to 

8 the compounding or carry-over effects on tree growth. 

9 Given the above, we conclude that the public welfare significance of the effects that can 

10 occur as a result of multiple year O3 exposures are greater than those associated with a single 

11 year. Thus, to the extent that the Administrator’s priority for public welfare protection to be 

12 afforded by the secondary O3 standard is on long-term effects that occur in sensitive tree species 

13 in natural forested ecosystems including federally protected areas such as Class I areas or on 

14 lands set aside by States, Tribes and public interest groups to provide similar benefits to the 

15 public welfare, a standard with a form that evaluates the cumulative seasonal index across 

16 multiple years might be considered to provide an appropriate match to the nature of O3-related 

17 effects on vegetation upon which the secondary O3 standard is focused. In considering such 

18 forms, we focus on one that averages the W126 index values across three years, as discussed in 

19 section 6.2 above. 

20 We take note, however, of comments from CASAC on this matter, in particular their 

21 comment in the current review that “...[a]veraging across years is not recommended because a 

22 single high exposure year could have lasting effects because of the perennial nature of many 

23 plants and the lag times associated with propagating effects through ecosystem trophic levels. 

24 Averaging would obscure such critical impacts and lead to inadequate protection against welfare 

25 effects” (Frey and Samet, 2012, p. 2). We recognize that annual effects on perennials can 

26 propagate into subsequent years, and thus first consider the available analyses of year-to-year 

27 variability in W126 index values. For example, based on an analysis of the inter-annual 

28 variability of seasonal W126 index values (using 2008-2010 data from the AQS database), the 

29 W126 index values can vary significantly from year to year (see Appendix 2C). This is not 

30 unexpected given the logistic weighting function and also inter-annual variability in 

31 meteorological conditions which contribute to O3 formation (see Appendix 2C). As a result, 

32 areas that meet a three-year average standard form for which there is substantial inter-annual 

33 variability of seasonal W126 index, are likely to have some years that are below the level of the 

34 standard and others that are at or above. Given this fact, several important implications should 

35 be noted. First, in regard to implications for potential long-term compounding of vegetation or 

36 ecosystem effects, it would be expected that annual impacts in years with cumulative air quality 
6-41 

1 below that of the three year average would allow species to do better than the target level of 

2 protection expected to be achieved on average across the three year period. Thus, staff note the 

3 importance, when considering an appropriate level for a form that averages W126 index values 

4 across three years, of considering the extent to which the cumulative effect of different average 

5 W126 exposures across the three-year period would be judged adverse. 

6 Additionally, in regard to implications for standard stability, a standard based on an 

7 annual W126 index would be expected to have a lower degree of year-to-year stability relative to 

8 a standard based on a form that averages seasonal indices across three consecutive years, given 

9 the potential for large year-to-year variability in annual W126 values in areas across the country. 

10 Thus, a three-year evaluation period can contribute to greater public welfare protection by 

11 limiting year-to-year disruptions in ongoing control programs that would occur if an area was 

12 frequently shifting in and out of attainment due to extreme year-to-year variations in 

13 meteorological conditions. This greater stability in air quality management programs thus 

14 facilitates achievement of the protection intended by a standard. 

15 Thus, to the extent that the Administrator puts greater weight on protecting those effects 

16 associated with multi-year exposures and given the described public welfare benefit of having a 

17 standard of a form with more year-to-year stability, we conclude that it is appropriate to consider 

18 a secondary standard form that averages the seasonal W126 index values across three 

19 consecutive years to achieve the appropriate target level of protection for longer-term effects, 

20 including compounding, and to achieve greater stability in air quality management programs, and 

21 thus, public welfare protection, than might result from an annual standard. 

22 Turning to consideration of an appropriate range of levels for a W126 based standard, we 

23 first note that the available O3-related vegetation effects evidence reflects a continuum from 

24 relatively higher O3 concentrations, at which scientists generally agree that vegetation effects are 

25 likely to occur, through lower concentrations at which the likelihood and magnitude of a 

26 response become increasingly uncertain. Further, we recognize the different types of O3-induced 

27 effects and genetic variability within and between species which contribute variability to 

28 observed responses across species. In light of this, we recognize the role of the Administrator’s 

29 judgments regarding the adversity of the known and anticipated effects to the public welfare. 

30 Thus, the EPA has developed a paradigm to assist the Administrator in putting the available 

31 science and exposure/risk information into the context of public welfare (as discussed in section 

32 5.1 above). This paradigm has evolved over the course of the O3 NAAQS reviews and has also 

33 been informed by similar constructs developed for other secondary NAAQS reviews. Most 

34 recently, this paradigm has expanded to include consideration of adversity in terms of effects on 

35 the ecosystem services associated with identified O3-induced effects (see discussion in 6.3 

36 above). 
6-42 

1 In considering the range of W126 index values most recently cited by CASAC, which 

2 generally correspond to conclusions from the 1996 workshop, we note that the tree seedling 

3 biomass loss percentages were per-year percentages, with consideration of longer-term impacts. 

4 Staff is mindful of this in light of our consideration of a potential alternative standard with a 

5 multi-year form. We additionally take note of the workshop objectives to identify values that 

6 might provide protection of vegetative components of natural ecosystems (recognizing foliar 

7 injury as the indicator), protection of some aspect of the integrity of forest ecosystems (using 

8 growth effects on seedlings as surrogate) and protection against crop yield reductions (while 

9 acknowledging uncertainties). In considering the range that derives, at least in part, from this 

10 workshop, we additionally recognize the need for our purposes with regard to a secondary O3

11 standard, to consider the public welfare significance of identified effects, as discussed in section 

12 6.3 above. We find that this need to consider public welfare significance may lead to 

13 identification of a somewhat different range of W126 index values as appropriate to consider for 

levels for a W126-based standard.15 14 

15 In considering potential levels for an alternative standard based on the W126 metric, we 

16 find it useful to consider the observations of biomass loss and crop yield loss in Table 6-1 above. 

17 In so doing, we take note of the different index value estimates with regard to number of studied 

18 species below different response benchmarks, as well as the median response. For example, we 

19 note that the tree seedling relative biomass loss estimates for 15-17 ppm-hrs include five of the 

20 twelve studied species below 2%, five to six species below 5% and nine to ten species below 

21 10%, as well as a median species response of 5 to 6%. At these index values, the median crop 

22 yield loss estimate across studied crops is just at or below 5% (and eight of ten are below 10%). 

23 At the lower end of the index values in Table 6-1, the tree seedling estimates for 7 ppm-hrs 

24 include ten species below 5%, seven below 2% and a median response just at or below 2%. We 

25 additionally consider the WREA estimates which indicate 143 or 144 of the 145 assessed Class I 

26 areas with tree seedling weighted relative biomass loss estimates below 2% for air quality 

27 scenarios representing W126 values of 15 and below. Such estimates are of particular relevance 

28 to judgments required of the Administrator regarding effects of public welfare importance. 

29 Further we note other WREA estimates indicating potential benefits for effects related to public 

30 welfare, such as carbon sequestration and pollutant removal. In so doing, however, we also take 

31 note of the appreciable uncertainty in these quantitative estimates, and the policy judgements 

32 required of the Administrator with regard to consideration of such uncertainties. 

33 On the basis of all the considerations described above, including the evidence and 

34 exposure/risk analyses, and advice from CASAC, we conclude that an appropriate range of 

 

15 We additionally recognize variability in derivation of W126 estimates generally equivalent to SUM06 

estimates from the 1996 workshop, as noted in section 6.3 above.
6-43 

1 W126 index values for the Administrator to consider in identifying a target degree of public 

2 welfare protection, extends from 7 ppm-hrs to 15 ppm-hrs or somewhat higher (as further 

3 described below). In so doing, we primarily consider the evidence- and exposure/risk-based 

4 information for cumulative seasonal O3 exposures represented by W126 index values (including 

5 those represented by the WREA average W126 scenarios) associated with biomass loss in 

6 studied tree species, both in and outside areas that have been afforded special protections. We 

7 additionally recognize foliar injury as an important O3 effect which, depending somewhat on 

8 severity and spatial extent, may reasonably be concluded to be of public welfare significance 

9 when occurring in nationally protected areas. However, we additionally take note of the 

10 appreciable variability in this endpoint, as summarized in chapter 5 and section 6.3 above, which 

11 poses challenges to giving it primary emphasis in identifying potential alternative standard 

12 levels. Similarly, we give less emphasis to consideration of crop yield loss in our consideration 

13 of potential standard levels here and in section 6.3 above, noting the median estimates of 

14 approximately 5% or lower for W126 index levels at and below 17 ppm-hrs. We also note the 

15 range of factors affecting annual crop yields, including those related to the role of management 

16 strategies as recognized in sections 5.3 and 6.2 above which complicate the identification of a 

17 degree of impact that can be considered adverse to the public welfare. On the other hand, tree 

18 biomass loss can be an indicator of more significant ecosystem-wide effects which might 

19 reasonably be concluded to be significant to public welfare. For example, when it occurs over 

20 multiple years at a sufficient magnitude, it is linked to an array of effects on other ecosystem21 level processes, such as nutrient and water cycles, changes in above and below ground 

22 communities, carbon storage and air pollution removal (U.S. EPA, 2014, Figure 5-1), that have 

23 the potential to be adverse to the public welfare. 

24 In focusing on trees and their associated ecosystem services, we first note that the studied 

25 tree species vary widely in their sensitivity to O3-induced relative biomass loss. For example, 

26 2% seedling biomass loss is estimated to occur with cumulative seasonal O3 exposure in terms of 

27 W126 index values below 10 ppm-hrs for seven of the studied species and at or above 

28 approximately 20 ppm-hrs in the other five studied species (Figure 5-1 above). The median 

29 W126 index value (across studied species) for which a 2% biomass loss is estimated ranges 

30 between approximately 7 and 14 ppm-hrs, among the three approaches presented in the WREA 

31 (see description in section 5.2.1). In considering the potential magnitude of the ecosystem 

32 impact of tree species, we focus on the WREA estimates of weighted relative biomass loss for 

33 the W126 air quality scenarios (U.S. EPA, 2014, section 6.8). For the current standard and the 

34 three W126 scenarios, these estimates indicate weighted relative biomass loss less than or equal 

35 to 2% in 143-144 of 145 assessed nationally protected Class I areas. To the extent that emphasis 

36 is given to such estimates for nationally protected Class I areas and for appreciable percentages 
6-44 

1 of forested areas nationwide, a W126 index value extending up to 15 ppm (or perhaps somewhat 

2 higher) may be appropriate to consider. 

3 In considering the evidence-based information regarding tree seedling growth effects and 

4 specifically biomass loss, we recognize an array of W126 index values that may be appropriate 

5 to consider depending on the weight placed on different policy-related objectives with regard to 

6 proportions of the 12 studied tree species at or below different growth response benchmarks and 

7 on associated uncertainties. For example, conclusions may be reached regarding the higher index 

8 values (e.g., up to 17 ppm-hrs) to the extent weight is placed on W126 index values for which 

9 tree seedling C-R functions project less than 2% biomass loss in approximately half of the 

10 studied species and less than 10% biomass loss in the large majority of studied species and on 

11 index values for which weighted RBL is below 2% in nearly all assessed Class I areas, and to the 

12 extent that estimates associated with the remaining species are judged too variable and/or 

13 uncertain. For index values of 15 or 17 ppm-hrs, the species-specific composite C-R functions 

14 indicate relative biomass loss less than 2% in at least five of the 12 studied species and less than 

15 10% in at least nine or ten, in addition to an estimated median across studied species of 

16 approximately 5-6% . Additionally, the WREA estimates for all three W126 scenarios include 

17 weighted RBL below 2% in 143 to 144 of 145 assessed Class I areas. Alternatively, to the 

18 extent weight is given to median biomass across tree species of no more than 2% based on the 

19 evidence-based C-R functions, and to the potential, while uncertain, for appreciable gains in 

20 carbon sequestration, a focus on the lower end of the range (e.g., down to 7 ppm-hr) may be 

21 appropriate. 

22 Thus, in staff’s view, the evidence- and exposure/risk-based information relevant to tree 

23 biomass loss and the associated ecosystem services important to the public welfare support 

24 consideration of a W126-based secondary standard with index values within the assessed range 

25 of 7-15 ppm-hrs or somewhat higher (e.g., 17 ppm-hrs). We consider such a range for a 

26 potential alternative cumulative seasonal W126-based standard, averaged over three years. In so 

27 doing, we take note of CASAC’s advice regarding the importance of considering the lasting or 

28 carry-over effects that can derive from single year exposures of perennial plants, and recognize 

29 the importance of considering the available evidence and exposure/risk based information related 

30 to such effects, as well as associated uncertainties. We additionally recognize uncertainty 

31 associated with any characterization of a relationship between the level of protection afforded for 

32 cumulative growth-related effects by potential alternative W126-based standards of single year 

33 or three-year average form. Lastly, we are mindful of the public welfare judgments required of 

34 the Administrator with regard to the public welfare significance of identified effects and the 

35 requisite level of protection, as well as the appropriate weight to assign the range of uncertainties 

36 inherent in the evidence and analyses.
6-45 

1 In also noting the potential ecosystem services benefits related to tree biomass loss for 

2 potential alternative W126-based standards in light of the qualitative and semi-quantitative 

3 information in the WREA regarding the types and potential magnitude of impacts on associated 

4 services, we recognize, in particular that impacts on climate can reasonably be concluded 

5 significant from a public welfare perspective and CO2 sequestration has been identified as a 

6 potentially important tool for managing anthropogenic impacts on climate. However, we 

7 additionally take note of significant uncertainties and limitations associated with WREA 

8 estimates related to carbon sequestration. Thus, in selecting a target level of protection for forest 

9 trees and their associated ecosystem services, the Administrator will need to exercise judgments 

10 regarding the appropriate weight to place on the potential for benefits to the public welfare with 

11 respect to ecosystem services of carbon storage and urban air pollution removal associated with 

12 tree growth, as well as the large uncertainties associated with this information. 

13 We further recognize that public welfare considerations taken into account by the 

14 Administrator may have the potential to affect the target protection judged requisite by the 

15 Administrator and the associated range of W126 index values. For example, to the extent the 

16 Administrator chooses to put more weight on effects associated with longer-term conditions, it 

17 may be appropriate to evaluate the significance of these longer term effects to the public welfare, 

18 as well as the role that year-to-year exposure variability can play in realizing the potential public 

19 welfare impacts. In so doing, the Administrator may put weight on a range of percentages of 

20 biomass loss (either greater or less than 2%) with an objective to achieve substantial protection 

21 of a large proportion of the studied species. Thus, in considering a range of 3-year average 

22 W126 index values appropriate to provide the target protection, we recognize that the 

23 Administrator may consider a range somewhat beyond the staff identified range. 

24 In considering the lower end of the staff-identified range, the Administrator would need 

25 to put more weight on the uncertainties that suggest the analyses may be underestimating the 

26 public welfare impacts associated with different cumulative exposure levels. These uncertainties 

27 can include the relatively small number of O3 sensitive species for which we have robust C-R 

28 functions included in the analyses, O3-sensitive species for which we have no C-R function, the 

29 lack of information regarding the relationship of O3 effects on tree seedlings (e.g., relative 

30 biomass loss) that occur in one or a few growing seasons to longer-term effects that might be 

31 expected to accrue over the life of the trees extending into adulthood, the paucity of ambient air 

32 monitoring data in some areas (e.g., the west) that leads to less than complete national coverage, 

33 recognition that co-occurring stressors may in some cases exacerbate predicted O3-induced 

34 effects, and the limited number of WREA urban area case studies, uncertainties associated with 

35 the model-adjusted air quality exposure surfaces,and the inability to quantify some potentially 
6-46 

1 important associated ecosystem services and incremental impacts (i.e. for insect damage, fire

2 regimes). 

3 In considering a range somewhat above that of the staff-identified range, the 

4 Administrator would need to put more weight on the uncertainties suggesting the potential for 

5 overestimating the beneficial public welfare impacts that might be achieved at 3-year average 

6 levels within the staff-identified range. These uncertainties include the lack of information 

7 (particularly quantitative) on the relationship of tree seedling O3 effects (e.g., relative biomass 

8 loss) that occur in one or a few growing seasons to longer-term effects that might be expected to 

9 accrue over the life of the trees extending into adulthood, recognition that co-occurring stressors 

10 or environmental factors (e.g., drought) may in some cases mitigate predicted O3-induced effects, 

11 the variability in number of C-R functions available for each of the 12 studied tree species, and 

12 information concerning the extent to different endpoints and effects might be considered adverse 

13 to public welfare . Given these and potentially other uncertainties, the Administrator may choose 

14 to select a higher range of levels judged to have less potential to provide overprotection. 

15 Lastly, we also conclude that, to the extent the Administrator finds it useful to consider 

16 the public welfare protection that might be afforded by a revised primary standard, this is 

17 appropriately judged through the use of a cumulative seasonal W126-based exposure metric. 

18 Such a use could inform a judgment of whether the primary standard would be expected to 

19 achieve the level of public welfare protection concluded to be requisite under the Act in terms of 

20 a metric considered appropriate to judging impacts on public welfare. See Mississippi. 723 F. 3d 

21 at 272-73. The staff further concludes that the drawing of conclusions with regard to the public 

22 welfare protection afforded by such a standard, as well as identification of the requisite level of 

23 protection for such a standard, should entail consideration of the air quality conditions likely to 

24 be achieved in terms of the cumulative seasonal W126-based metric described above. In such a 

25 consideration, such as through the review of overlap analyses discussed in section 6.2 above, 

26 staff further concludes it is important to take into account associated uncertainties, including 

27 those associated with the limited monitor coverage in many rural areas, including those in the 

28 west and southwest and at high elevation sites. 

29 6.6 SUMMARY OF PRELIMINARY CONCLUSIONS ON THE SECONDARY 

30 STANDARD

31 Staff preliminary conclusions are informed by our consideration of the available 

32 scientific evidence as assessed in the ISA, the air quality/exposure/risk information in the second 

33 draft WREA, advice from CASAC thus far in this review and in prior reviews, and public 

34 comment thus far in this review. Staff conclusions in the final PA will be further informed by 

35 comments from CASAC and the public on this draft document and by the final WREA. 
6-47 

1 Staff preliminary conclusions on policy options that are appropriate for the 

2 Administrator’s consideration in making decisions on the secondary standards for O3, together 

3 with supporting conclusions from sections 5.7 and 6.5 above, are briefly summarized below. In 

4 reaching conclusions on alternative standards to provide requisite protection for public welfare

5 effects associated with ambient O3 exposures, staff has considered these standards in terms of the 

6 basic elements of the NAAQS: indicator, form, averaging time, and level. In drawing these 

7 conclusions, we are mindful that the Act requires secondary standards to be set so that, in the 

8 Administrator’s judgment, they are requisite to protect public welfare from known or anticipated 

9 adverse environmental effects, such that the standards are to be neither more nor less stringent 

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

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

12 (1) Staff preliminarily concludes, based on the combined consideration of the body of 

13 evidence and the results from the quantitative exposure/risk assessment, that the 

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

15 of the public welfare protection provided by the current standard and it is 

16 appropriate to consider revising the standard to provide greater public welfare 

17 protection. 

18 (2) In considering an appropriate target level of protection for a revised standard, 

19 staff additionally preliminarily concludes that it is appropriate to judge O3 public 

20 welfare impacts using the cumulative seasonal W126-based metric.

21 a. To the extent the Administrator finds it useful to consider the extent of 

22 public welfare protection that might be afforded by a revised primary 

23 standard, staff preliminarily concludes that public welfare protection is 

24 appropriately judged through the use of the cumulative seasonal W126-

25 based metric.

26 (3) With regard to indicator, staff preliminarily concludes that it is appropriate to 

27 continue to use O3 as the indicator for a standard that is intended to address 

28 welfare effects associated with exposure to O3, alone or in combination with 

29 related photochemical oxidants. Based on the available information, staff 

30 preliminarily concludes that there is no basis for considering an alternative 

31 indicator at this time.

32 (4) With regard to averaging time and form, staff preliminarily concludes that it is 

33 appropriate to consider a revised secondary standard in terms of the cumulative, 

34 seasonal, concentration-weighted form, the W126 index. With regard to 
6-48 

1 definition of the W126 index for this purpose, staff makes the additional 

2 preliminary conclusions. 

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

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

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

6 period generally coincides with maximum biological activity for most 

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

8 growing seasons. 

9 b. It is appropriate to cumulate daily exposures for the 12-hr period from 8 

10 am to 8 pm, generally representing the daylight period during the 3-month 

11 period identified above. 

12 c. It is appropriate to consider a form that averages W126 index values 

13 across three consecutive years. Staff concludes it is appropriate to 

14 consider this form in conjunction with appropriate levels in order to 

15 provide the desired degree of public welfare protection from O3 effects 

16 across multiple years. 

17 With regard to level for a standard as described above, we preliminarily conclude that it 

18 is appropriate to give consideration to a range of levels from somewhat above 15 ppm-hrs to 7 

19 ppm-hrs, expressed in terms of the W126 index. Staff additionally notes that, consideration of 

20 the support provided by the information available in this review will depend on public welfare 

21 policy judgments by the Administrator regarding the protection of public welfare. This range 

22 reflects staff judgment that a standard set within this range could provide an appropriate degree 

23 of public welfare protection. 

24 6.7 KEY UNCERTAINTIES AND AREAS FOR FUTURE RESEARCH AND 

25 DATA COLLECTION

26 Staff believes it is important to highlight key uncertainties associated with establishing 

27 secondary standards for O3. Such key uncertainties and recommendations for welfare-related 

28 research, model development, and data gathering are outlined below. In some cases, research in 

29 these areas can go beyond aiding standard setting to aiding in the development of more efficient 

30 and effective control strategies. We note, however, that a full set of research recommendations 

31 to meet standards implementation and strategy development needs is beyond the scope of this 

32 discussion. Based on items highlighted in chapter 9 of the ISA and chapter 5 and 6 herein,we 

33 have identified the following key uncertainties, research questions and data gaps that have been 

34 highlighted in this review of the welfare-based secondary standard. The first set of key 
6-49 

1 uncertainties and research recommendations discussed below is that associated with the 

2 extrapolation to plant species and environments outside of specific experimental or field study 

3 conditions. The second set of key uncertainties and research recommendations pertain to our 

4 ability to assess the impact of O3 on other welfare effects categories such as climate, ecosystem 

5 components such as wildlife, and whole ecosystem structure and function. Third, we identify 

6 research areas related to the development of approaches, tools, or methodologies useful in 

7 characterizing the relationship between O3 and public welfare in a policy context. These three 

8 areas are described below.

9 There have been five decades of research regarding O3 effects on plants and much 

10 information has been compiled in previous reviews. One of the most important research results 

11 for the review of the secondary O3 standard is C-R relationships for plant species. This review 

12 uses C-R functions from 22 crop and tree species. However, there are tens of thousands of plant 

species in the U.S. (USDA, NRCS. 201416 13 ) and 66 plant species have been identified as O3

sensitive on National Park Service and US Fish and Wildlife Service lands17 14 . Studies using large 

15 numbers of native plant species across regions where those species are indigenous, might be 

16 expected to reduce uncertainties associated with extrapolating plant response for a given level of 

17 O3 using composite response functions across differing regions and climates. Research on 

18 additional species might additionally improve our understanding of the full range of response of 

19 plant species to O3. Studies focused on fruits and vegetables might assist in reducing 

20 uncertainties associated with O3 effects on agriculture. Particular focus is suggested on 

21 organically grown vegetables that may receive less intensive management than conventionally 

22 grown crops. Recent studies indicate that watermelons may be particularly sensitive to O3

23 exposure (U.S. EPA, 2013, section 9.4.4.1) and older studies indicate grapes, honeydew melon, 

24 lemons and oranges may also be O3 sensitive (Abt Associates Inc., 1995). 

25 National visible foliar injury surveys can indicate how widespread O3 effects may be 

26 within the US. However, there remain uncertainties about the nature of the effects indicated by 

27 the observed foliar injury. These include uncertainty associated with estimating the risk to 

28 vegetation of differing amounts of O3-induced visible foliar injury over the plant’s leaf area and 

29 the relationship between relative soil moisture and the incidence and severity of foliar injury in 

30 sensitive species, as well as the extent to which visible foliar injury impacts ecosystem services 

31 (e.g., tourism). 

 

16 USDA, NRCS. 2014. The PLANTS Database (http://plants.usda.gov, 3 January 2014). National Plant 

Data Team, Greensboro, NC 27401-4901 USA.

17See http://www2.nature.nps.gov/air/Pubs/pdf/flag/NPSozonesensppFLAG06.pdf

 
6-50 

1 Some new information has emerged linking effects on tree seedlings with larger trees and 

2 similarities in results between exposure techniques (U.S. EPA 2013, section 9.6). Uncertainties 

3 remain in this area and in relationships between effects on individual plants and ecosystem 

4 effects. There are also uncertainties in extrapolating from O3 effects on juvenile to mature trees 

5 and from trees grown in the open versus those in a closed forest canopy in a competitive 

6 environment. Uncertainties in extrapolating individual plant response spatially or to higher 

7 levels of biological organization, including ecosystems, could be informed by research that 

8 explores and better quantifies the nature of the relationship between O3, plant response and 

9 multiple biotic and abiotic stressors, including those associated with climate change. Because 

10 these uncertainties are multiple and significant due to the complex interactions involved, new 

11 research will likely require a combination of manipulative experiments with model ecosystems, 

12 community and ecosystem studies along natural O3 gradients, and extensive modeling efforts to 

13 project landscape-level, regional, national and international impacts of O3. 

14 Uncertainties associated with projections of the effects of O3 on the ecosystem processes 

15 of water, carbon, and nutrient cycling, particularly at the stand and community levels might be 

16 addressed through research on the effects on below ground ecosystem processes in response to 

17 O3 exposure alone and in combination with other stressors. These below-ground processes 

18 include interactions of roots with the soil or microorganisms, effects of O3 on structural or 

19 functional components of soil food webs and potential impacts on plant species diversity, 

20 changes in the water use of sensitive trees, and if the sensitive tree species is dominant, potential 

21 changes to the hydrologic cycle at the watershed and landscape level. Research on competitive 

22 interactions under different O3 exposures might improve our understanding of how O3 may 

23 affects biodiversity or genetic diversity. Such research could be strengthened by modern 

24 molecular methods to quantify impacts on diversity. 

25 Important interactions with biotic and abiotic stressors have been identified in this 

26 review. More tools and research would improve our understanding of relationships between O3

27 exposure and stressors such as insect infestations, plant diseases, drought and potential stressors 

28 from climate change. It is also important to understand how such interactions may affect 

29 ecosystem services such as CO2 sequestration; food and fiber production; wildlife habitat and 

30 water resources.

31 One of the most important uncertainties in this review is the characterization of air 

32 quality in rural areas where there is limited monitoring. More comprehensive monitoring in 

33 these areas would reduce uncertainties associated with O3 exposures in many rural areas. Areas 

34 of particular uncertainty include protected natural areas in the western U.S, including those at 

35 high elevation, as well as those downwind of recently expanded oil and gas development areas. 

36 Uncertainties associated with quantifying exposure in areas with and without monitors might be 
6-51 

1 addressed through additional work on interpolation methods and air quality models that are 

2 tailored to estimating cumulative seasonal exposures, as well as on more refined spatial grids and 

3 complex terrain. 

4 6.8 REFERENCES

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6 Federal Register 1996. National Ambient Air Quality Standards for Ozone; Proposed Rule. 40 CFR 50; Federal 

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14 Federal Register 2010. National Ambient Air Quality Standards for Ozone; Proposed Rule. 40 CFR 50 and 58; 

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25 Henderson, R. (2006) Letter from CASAC Chairman Rogene Henderson to EPA Administrator Stephen Johnson. 

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29 Henderson, R. (2008) Letter from CASAC Chairman Rogene Henderson to EPA Administrator Stephen Johnson. 

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31 Samet, J.M. (2010) Clean Air Scientific Advisory Committee (CASAC) Review of EPA’s Proposed Ozone National 

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35 USDA, NRCS. 2014. The PLANTS Database (http://plants.usda.gov, 3 January 2014). National Plant Data Team, 

36 Greensboro, NC 27401-4901 USA
6-52 

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16 Wolff, G.T. (1996) Letter to EPA Administrator Carol Browner, RE: Closure by the Clean Air Scientific Advisory 

17 Committee (CASAC) on the Secondary Standard Portion of the Staff Paper for Ozone. EPA-SAB-CASAC18 LTR-96-002, April 4, 1996. 
January 2013 Draft – Do Not Quote or Cite 

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: Annual and 3-Year Average 

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

 Appendix 3A. Modes of Action Summary .............................................................................. 3A-1 

 Appendix 3B. Recent Studies of Respiratory-related Emergency Department Visits 

 and Hospital Admissions ................................................................................. 3B-1 

Hospital Admissions for All Respiratory Causes ........................................ 3B-1

Cause-Specific Hospital Admissions ........................................................... 3B-5

Emergency Department Visits for All Respiratory Causes ......................... 3B-7

Cause-Specific Emergency Department Visits ............................................ 3B-7

 Appendix 3C. At-risk Populations. .......................................................................................... 3C-1

 Appendix 3D. Air Quality Data for Locations of Key Epidemiological Studies ................... 3D-1 

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

 Appendix 6A. Calculation of Approximate Equivalent 12-hr SUM06 and 12-hr W126 ....... 6A-1 
2A-1 

1 APPENDIX 2A 

2 SUPPLEMENTAL AIR QUALITY MODELING ANALYSES OF 

3 BACKGROUND OZONE

 4 

5 Table of Contents

6 List of Figures ............................................................................................................................ 2 

7 List of Tables .............................................................................................................................. 4 

8 1. Introduction .................................................................................................................. 5 

9 2. Description of modeling methodologies .................................................................. 6 

10 a. 2007 GEOS-Chem/CMAQ zero-out modeling ........................................................... 7 

11 b. 2007 GEOS-Chem/CAMx source apportionment modeling .............................. 12

12 3. Estimates of seasonal-average background ozone levels ....................................... 15

13 4. Distributions of background ozone levels .................................................................. 22

14 5. Contribution of various processes and sources to total background ozone ...... 30

15 6. Estimates of the fractional background contribution to total ozone in 12 specific 

16 areas .............................................................................................................................. 37

17 7. Background ozone and W126 ..................................................................................... 39

18 8. Summary ........................................................................................................................ 39

19 9. References ..................................................................................................................... 42

20
2A-2 

21 List of Figures

22 Figure 1a. Modeling domain used in 2007 CMAQ and CAMx modeling ........................................... 10

23

24 Figure 1b. Density scatterplot comparing CMAQ base daily peak 8-hour ozone predictions 

25 against observed 8-hour ozone peaks paired in space and time for all sites during April26 October 2007 ................................................................................................................................................ 10

27

28 Figure 1c. Bias in seasonal mean (April-October) maximum daily 8-hour ozone predictions in 

29 the 2007 CMAQ base simulation ...............................................................................................................11

30

31 Figure 1d. Relationship between CMAQ estimations of MDA8 natural background ozone and 

32 daily model biases .........................................................................................................................................11

33 

34 Figure 2a. Density scatterplot comparing CAMx base daily peak 8-hour ozone predictions 

35 against observed 8-hour ozone peaks paired in space and time for all sites during April-October 

36 2007 ................................................................................................................................................................ 14 

37

38 Figure 2b. Bias in seasonal mean (April-October) maximum daily 8-hour ozone predictions in 

39 the 2007 CAMx base simulation .................................................................................................................14

40

41 Figure 3a. April-October average MDA8 ozone (ppb) at monitoring locations across the U.S. as 

42 estimated by a 2007 CMAQ base simulation ............................................................................................18

43

44 Figure 3b. April-October average natural background MDA8 ozone (ppb) at monitoring 

45 locations across the U.S. as estimated by a 2007 CMAQ zero out simulation .................................... 18

46

47 Figure 3c. April-October average North American background MDA8 ozone (ppb) at monitoring 

48 locations across the U.S. as estimated by a 2007 CMAQ zero out simulation .....................................19

49

50 Figure 3d. April-October average United States background MDA8 ozone (ppb) at monitoring 

51 locations across the U.S. as estimated by a 2007 CMAQ zero out simulation .....................................19

52

53 Figure 4a. Ratio of natural background to total April-October average MDA8 ozone at 

54 monitoring locations across the U.S. as estimated based on 2007 CMAQ simulations ..................... 20

55

56 Figure 4b. Ratio of N. American background to total April-October average MDA8 ozone at 

57 monitoring locations across the U.S. as estimated based on 2007 CMAQ simulations ..................... 20

58

59 Figure 4c. Ratio of U.S. background to total April-October average MDA8 ozone at 

60 monitoring locations across the U.S. as estimated based on 2007 CMAQ simulations ..................... 21

61

62 Figure 4d. Ratio of sources other than U.S. anthropogenic emissions to total April-October 

63 average MDA8 ozone at monitoring locations across the U.S. as estimated by a 2007 CAMx 

64 source apportionment simulation ..............................................................................................................21 

65

66 Figure 5a. Distribution of natural background MDA8 ozone (ppb) at monitoring locations 

67 across the U.S. (Apr-Oct), binned by base modeled site-day MDA8, as estimated by 2007 
2A-3 

68 CMAQ simulations ..................................................................................................................................... 25

69

70 Figure 5b. Distribution of N. American background MDA8 ozone (ppb) at monitoring 

71 locations across the U.S. (Apr-Oct), binned by base modeled site-day MDA8, as estimated by 

72 2007 CMAQ simulations ............................................................................................................................ 25

73

74 Figure 5c. Distribution of U.S. background MDA8 ozone (ppb) at monitoring locations across 

75 the U.S. (Apr-Oct), binned by base modeled site-day MDA8, as estimated by 2007 CMAQ 

76 simulations .................................................................................................................................................. 26

77

78 Figure 5d. Distribution of MDA8 ozone contributions from non-U.S. manmade sources (ppb) 

79 at monitoring locations across the U.S. (Apr-Oct), binned by base modeled site-day MDA8, as 

80 estimated by 2007 CAMx simulations ..................................................................................................... 26

81

82 Figure 6a. Distribution of natural background MDA8 ozone fractions at monitoring locations 

83 across the U.S. (Apr-Oct), binned by base modeled site-day MDA8, as estimated by 2007 

84 CMAQ simulations ...................................................................................................................................... 27

85

86 Figure 6b. Distribution of N. American background MDA8 ozone fractions at monitoring 

87 locations across the U.S. (Apr-Oct), binned by base modeled site-day MDA8, as estimated by 

88 2007 CMAQ simulations ............................................................................................................................. 27

89

90 Figure 6c. Distribution of U.S. background MDA8 ozone fractions at monitoring locations 

91 across the U.S. (Apr-Oct), binned by base modeled site-day MDA8, as estimated by 2007 

92 CMAQ simulations ....................................................................................................................................... 28

93

94 Figure 6d. Distribution of MDA8 ozone fractions from non-U.S. anthropogenic sources at 

95 monitoring locations across the U.S. (Apr-Oct), binned by base modeled site-day MDA8, as 

96 estimated by the 2007 CAMx simulation .................................................................................................. 28

97

Figure 7. April-October 95th 98 percentile United States background MDA8 ozone (ppb) at 

99 monitoring locations across the U.S. as estimated by a 2007 CMAQ base simulation ..................... 29

100

101 Figure 8a. Difference in April-October average MDA8 ozone (ppb) at monitoring locations 

102 across the U.S. between the USB scenario and the NAB scenario ...................................................... 32

103

104 Figure 8b. Difference in April-October average MDA8 ozone (ppb) at monitoring locations 

105 across the U.S. between the NAB scenario and the NB scenario .......................................................... 32 

106

107 Figure 9a. Percentage of April-October average MDA8 ozone that is apportioned to 

108 boundary conditions as estimated at monitoring locations by a 2007 CAMx simulation ................. 33

109

110 Figure 9b. Percentage of April-October average MDA8 ozone that is apportioned to U.S.

111 anthropogenic sources as estimated at monitoring locations by a 2007 CAMx simulation ............ 33

112

113 Figure 9c. Percentage of April-October average MDA8 ozone that is apportioned to purely 

114 biogenic emissions as estimated at monitoring locations by a 2007 CAMx simulation .................... 34

115
2A-4 

116 Figure 9d. Percentage of April-October average MDA8 ozone that is apportioned to 

117 climatological fire emissions as estimated at monitoring locations by a 2007 CAMx simulation ... 34

118

119 Figure 9e. Percentage of April-October average MDA8 ozone that is apportioned to 

120 anthropogenic emissions from in-domain Canadian and Mexican sources as estimated at 

121 monitoring locations by a 2007 CAMx simulation ....................................................................................35

122

123 Figure 9f. Percentage of April-October average MDA8 ozone that is apportioned to Category 3 

124 marine vessel emissions beyond U.S. territorial waters as estimated at monitoring locations by 

125 a 2007 CAMx simulation ..............................................................................................................................35

126

127 Figure 9g. Percentage of April-October average MDA8 ozone that is apportioned to Gulf of 

128 Mexico point sources as estimated at monitoring locations by a 2007 CAMx simulation ............... 36

129

130

131 List of Tables

132

133 Table 1a. April-October average MDA8 ozone, average MDA8 ozone from sources other than 

134 U.S. manmade emissions, and the fractional contribution of these background sources in the 

135 12 REA urban study areas, as estimated by a 2007 CAMx simulation................................................. 37

136

137 Table 1b. Average MDA8 ozone, average MDA8 ozone from sources other than U.S. 

138 manmade emissions, and the fractional contribution of these background sources in the 12 REA 

139 areas, as estimated by a 2007 CAMx simulation using site-days in which base MDA8 ozone 

140 exceeded 60 ppb .......................................................................................................................................... 38

141

142 Table 1c. Fractional contribution of non-U.S. manmade emissions sources in the 12 REA urban 

143 study areas, as estimated by a 2007 CAMx simulation using means and medians of daily MDA8 

144 fractions ......................................................................................................................................................... 38

145

146 Table 1d. April-October average MDA8 ozone, average MDA8 ozone from USB, and the 

147 fractional contribution of these background sources in the 12 REA urban study areas, as 

148 estimated by two separate 2007 CMAQ simulations .............................................................................. 38

149
2A-5 

150 1. Introduction

151 One of the aspects of ozone that is unusual relative to the other pollutants with National 

152 Ambient Air Quality Standards (NAAQS) is that, periodically, in some locations, an appreciable fraction of 

153 the observed ozone results from sources or processes other than local and regional anthropogenic 

154 emissions of ozone precursors (Fiore et al., 2002). Any ozone formed by processes other than the 

155 chemical conversion of local or regional ozone precursor emissions, such as nitrogen oxides (NOx) or 

156 volatile organic emissions (VOC), is generically referred to as “background” ozone. As part of this review 

157 of the ozone NAAQS, EPA completed an extensive review of the known aspects of background ozone 

158 and summarized the findings in the Integrated Science Assessment (ISA) in March 2013 (USEPA, 2013). 

159 The purpose of this appendix is to present the results from supplemental air quality modeling analyses 

160 related to background ozone that were completed by EPA subsequent to the ISA. While these updated 

161 analyses use a recent base year (2007) and consider an alternative modeling methodology which can 

162 better account for non-linear ozone chemistry in some conditions, the results are largely consistent with

163 previous determinations about the magnitude of background ozone contributions across the U.S. 

164 Away from the surface, ozone can have an atmospheric lifetime on the order of weeks. As a 

165 result, background ozone can be transported long distances at heights above the boundary layer and, 

166 when meteorological conditions are favorable, be available to mix down to the surface and add to the 

167 total ozone loading from non-background sources. Generically, background ozone can originate from 

168 natural sources of ozone and ozone precursors, as well as from far upwind manmade emissions of ozone 

169 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 170 . Another important 

171 natural component of background is ozone that is naturally formed in the stratosphere through 

172 interactions of UV light with atomic oxygen (O2). Stratospheric ozone can periodically mix down to the 

173 surface at high concentrations, especially at higher altitude locations. The manmade portion of the 

174 background includes any ozone formed due to anthropogenic sources of ozone precursors emitted far 

175 away from the local area (e.g., international emissions). Finally, both biogenic and international 

176 anthropogenic emissions of methane, which can be chemically converted to ozone over relatively long 

177 time scales, can also contribute to global background ozone levels. 

178 The precise definition of background ozone can vary depending upon context, but it generally 

179 refers to ozone that is formed by sources or processes that cannot be influenced by actions within the 

180 jurisdiction of concern. In the first draft policy assessment document (EPA, 2012), EPA presented three 

181 specific definitions of background ozone: natural background, North American background, and U.S. 

182 background. Natural background (NB) was the narrowest definition of background and it was defined as 

183 the ozone that would exist in the absence of any manmade ozone precursor emissions. The other two 

184 previously-established definitions of background presume that the U.S. has little influence over 

185 anthropogenic emissions outside our continental or domestic borders. North American background 

186 (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 

187 emissions inside of North America. U.S. background (USB) is defined as that ozone that would exist in 

188 the absence of any manmade emissions inside the United States. It is important to note that each of 

189 these three definitions of background ozone requires photochemical modeling simulations to estimate 

190 what the residual ozone concentrations would be were the various anthropogenic emissions to be 

191 removed. 

192 As noted in the first draft policy assessment, EPA has revised several aspects of our 

193 methodology for estimating the change in health risk and exposure that would result from a revision to 

194 the ozone NAAQS. First, risk estimates are now based on total ozone concentrations as opposed 

195 previous reviews which only considered risk above background levels. Second, EPA is now using air 

196 quality models to estimate the spatial patterns of ozone that would result from attaining various levels 

197 of the NAAQS, as opposed to simplistic rollback techniques that required the estimation of a background 

198 ozone “floor” beyond which the rollback would not take place. Both of these revisions have had the 

199 indirect effect of obviating the need for estimating background ozone levels as part of the ozone risk 

200 and exposure assessment (REA). Regardless, EPA expects that a well-founded understanding of the 

201 fractional contribution of background sources and processes to surface ozone levels will be valuable 

202 towards informing policy decisions about the NAAQS. Section 2 of this document will describe the 

203 supplemental air quality modeling simulations that have recently been completed by EPA to bolster our 

204 understanding of background ozone. Section 3 will present the results from the updated analyses and 

205 provide estimates of average background ozone levels, and how they can vary in time and space across 

206 the U.S. Based on the same modeling, Section 4 will consider the entire spectrum of variable 

207 background ozone levels with special emphasis on areas and times in which background can approach or 

208 exceed the level of the NAAQS. Section 5 will utilize the supplemental air quality modeling estimates to 

209 determine the relative importance of specific components of background ozone. Section 6 will present 

210 estimates of the overall fraction of ozone that is estimated to result from background sources or 

211 processes in each of the 12 urban case study areas in the epidemiology study based analyses in Chapter 

212 7 of the Risk and Exposure Assessment (REA) (EPA, 2014) based on the updated modeling. Finally, 

213 Section 7 will conclude with a limited analysis of how background ozone levels impact longer-term 

214 ozone metrics that may be important from a welfare perspective (i.e., W126).

215 2. Description of modeling methodologies

216 As noted above, air quality models are typically used to estimate background ozone as it is quite 

217 difficult to measure directly. Without special monitoring, it is impossible to determine how much of the 

218 ozone measured by a monitor originated from sources that are considered background. Even the most 

219 remote monitors within the U.S. can periodically be affected by U.S. anthropogenic emissions. Previous 

220 modeling studies have estimated what background levels would be in the absence of certain sets of 

221 emissions by simply comparing the ozone differences between a base model simulation and a control 

222 simulation in which emissions were removed. This basic approach is often referred to as “zero out” 

223 modeling or “emissions perturbation” modeling. Examples of zero out modeling include the three major 

224 studies summarized in the ISA (Zhang et al., 2011; Emery et al., 2012, Lin et al., 2013). It is important to 

225 note that the specific concepts of NB, NAB, and USB are all explicitly tied to zero-out modeling, as those 

226 definitions are based on estimating what remains in the absence of specific sets of man-made emissions. 
2A-7 

227 EPA has conducted and will describe updated air quality modeling for a 2007 base year that employs a 

228 regional air quality model nested within a coarser-scale global chemical transport model to estimate NB, 

229 NAB, and USB levels when the respective manmade emissions are zeroed. This modeling is described in 

230 detail in section 2a. 

231 While the zero-out approach has traditionally been used to estimate background ozone levels, 

232 the methodology has some acknowledged limitations. First, from a policy perspective, the purely 

233 hypothetical and ultimately unrealizable zero manmade emissions scenarios have limited application in 

234 this regard. Secondly, the assumption that background ozone is what is left after specific emissions have 

235 been removed within the model simulation can be misleading in locations where ozone chemistry is

236 highly non-linear. Depending upon the local composition of ozone precursors, NOx emissions 

237 reductions can either increase or decrease ozone levels in the immediate vicinity of those reductions. 

238 For those specific urban areas in which NOx titration of ozone can be significant, zero-out modeling can 

239 result in inflated estimates of background ozone when these NOx emissions are completely and 

240 unrealistically removed. Paradoxically, in certain times and locations in a zero-out scenario there can be 

241 more background ozone than actual ozone within the model (EPA, 2014). 

242 A separate modeling technique attempts to circumvent these limitations by apportioning the 

243 total ozone within the model to its contributing source terms. This basic approach is referred to as 

244 “source apportionment” modeling. While source apportionment modeling has not been previously used 

245 in the context of estimating background ozone levels as part of an ozone NAAQS review, it has 

246 frequently been used in other regulatory settings to estimate the “contribution” to ozone of certain sets 

247 of emissions (EPA 2005, EPA 2011). The source apportionment technique provides a means of 

248 estimating the contributions of user-identified source categories to ozone formation in a single model 

249 simulation. This is achieved by using multiple tracer species to track the fate of ozone precursor 

250 emissions (VOC and NOx) and the ozone formation caused by these emissions. The methodology is 

251 designed so that all ozone and precursor concentrations are attributed to the selected source categories 

252 at all times without perturbing the inherent chemistry. The zero out modeling attempts to determine 

253 what ozone be in the absence of background sources. The source apportionment modeling attempts to 

254 determine how much of the modeled ozone has resulted from background sources. EPA has conducted 

255 and will describe new source apportionment modeling that employs a regional air quality model nested 

256 within a coarser-scale global chemical transport model to assess the contributions of boundary 

257 conditions and other potential background sources (e.g., wildfires, biogenic emissions, and 

258 Canadian/Mexican emissions). This modeling is described in detail in section 2b. 

259 a. 2007 GEOS-Chem/CMAQ zero-out modeling: 

260 In order to provide estimates of the overall fraction of ozone that is estimated to result from 

261 background sources in each of the 12 REA urban study areas, EPA conducted new modeling that utilized 

262 the same model base year (2007) as was used in the ozone modeling that inform the risk and exposure 

263 analyses (EPA, 2014, Appendix 4b). The EPA modeling used a model configuration similar to that of 

264 Emery (2012), in that it nested a regional-scale (12 km) air quality model inside a global air quality model
2A-8 

265 simulation with a much coarser horizontal grid resolution (2.0 by 2.5 degrees). Figure 1a shows a map of 

266 the model domain.

267 The global scale simulation utilized the GEOS-Chem model, version v8-03-02, except for the 

268 chemistry package which was from version v8-02-01. The emissions estimates used in the 2007 base 

269 year modeling were aggregated from a variety of sources, starting with the global Emissions Database 

270 for Global Atmospheric Research (EDGAR) emission inventory. These initial estimates were then 

271 improved by utilizing various area-specific inventories, such as the 2005 National Emissions Inventory 

272 (NEI) for the U.S. portions of the domain, and available inventories for Asia, Canada, Europe, and 

273 Mexico. In addition to the anthropogenic estimates, emissions were specified for a variety of 

274 background sources including: lightning NO, soil NOx, wildfires, and biogenic VOC emissions. The 

275 wildfire data is from the Global Fire Emissions Database (GFED). The biogenic VOC estimates were 

276 simulated by the Model of Emissions of Gases and Aerosols from Nature (MEGAN) version 2.1. The 

277 meteorological data is based on the Goddard Earth Observing System Model, Version 5 (GEOS-5) 

278 analysis fields. More information on the global simulation is available within Henderson et al. (2013). 

279 This reference also provides a broad evaluation of the ability of this specific GEOS-Chem configuration to 

280 provide accurate lateral boundary conditions of ozone to finer-scale regional simulations. Using satellite 

281 retrievals from the Tropospheric Emissions Spectrometer (TES), Henderson et al. (2013) concluded that 

282 the GEOS-Chem ozone prediction biases and errors are generally within TES uncertainty estimates. For 

283 instance, for the ozone season month of August, model predictions are within plus or minus 20 percent 

284 of the satellite estimates between nearly 80 percent of the time, with slightly better performance along 

285 the southern boundary.

286 The lateral boundary conditions from the global model were then used as inputs for a 12 km 

287 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 288 draft ozone 

289 REA and is described in more detail in Appendix 4b of EPA (2014), 2) a natural background run with 

anthropogenic ozone precursor emissions2 290 removed in both the global and regional models, 3) a North 

291 American background run with anthropogenic ozone precursor emissions removed across North 

292 America (global and regional model simulations), and 4) a U.S. background run with anthropogenic 

293 ozone precursor emissions were removed over the U.S (global and regional model simulations). 

294 Detailed analyses of EPA’s 2007 zero out modeling results are provided in sections 3 through 6 of this 

295 appendix. 

296 An operational model performance evaluation was completed for surface ozone in the 2007 

297 base simulation as described separately (EPA 2014, Appendix 4b). For the purposes of this analysis, EPA 

298 assessed the model ability to reproduce measured daily maximum 8-hour (MDA8) ozone values and 

299 seasonal mean MDA8 ozone concentrations for the period April to October 2007. As noted earlier, the 

300 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 

301 intended to capture discrete events from specific fires that occurred in 2007, so perfect correlation 

302 between observations and model predictions should not be expected. Figure 1b provides a density 

303 scatterplot of the observed and predicted daily 8-hour ozone peaks paired in space and time for the 

304 2007 CMAQ base. As can be seen, the majority of pairs line up along the 1:1 line. There is a tendency 

305 for the model to overestimate site-days with low 8-hour ozone peaks, and underestimate the site-days 

306 with higher peak ozone values. Modeled 8-hour ozone peak concentrations exhibited relatively small 

307 bias and error compared to the observations. The average bias in MDA8 ozone estimates was 3.5 ppb. 

308 Figure 1c depicts the spatial bias patterns in MDA8 ozone at all sites that measured valid ozone data for 

309 at least 100 days during the April-October period. CMAQ overestimations are greatest along the Gulf 

310 Coast region, along the Atlantic coastline, and over the central U.S. The majority of underestimated 

311 seasonal mean MDA8 occurs within southern California. The model performance for the 2007 base 

312 simulation is equivalent or better than typical state-of-the-science photochemical model performance 

313 recently reported in the literature (Simon et al, 2012).

314 Certainly some remote monitoring locations are more affected by background sources than 

315 other locations in the network. However, this and numerous other analyses have shown that even the 

316 most remote ozone monitoring locations in the U.S. are periodically affected by U.S. manmade 

317 emissions. In this analysis we carefully assess model performance to ensure that model error does not 

318 influence the characterization of background ozone. As noted in the recent ISA (EPA, 2013), there is

319 greater confidence in the ability of the model to predict mean contributions from background sources 

320 rather than individual events. Beyond the statistical analyses summarized in the previous paragraph and 

in appendix 4b of the 2nd 321 draft ozone REA (EPA, 2014), it is valuable to attempt to diagnose the model 

322 ability to account for background ozone within the simulation. EPA assessed whether any correlation 

323 existed between daily model biases and daily background ozone estimates. Figure 1d shows that at 

324 high-elevation sites (i.e., sites more than 1km above sea level) the highest estimates of natural 

325 background ozone tend occur on days with greatest overestimation. Conversely, the site-days with the 

326 lowest natural background estimates tend to occur when the model underestimates the observed daily 

327 peaks at these sites. This relationship between background estimates and simulation bias appears to be 

328 constrained to the mountainous portion of the Western U.S. Figure 1d also shows that estimates of 

329 natural background ozone greater than 60 ppb are associated with large over-predictions. However, 

330 based on the relatively low model bias and the general lack of correlation between daily bias values and 

331 background estimates, EPA believes that these model estimates can be used to help characterize 

332 background ozone levels over the U.S. Although the highest background estimates should be 

333 considered with caution. 
2A-10 

334 

335 Figure 1a. Modeling domain used in 2007 CMAQ and CAMx modeling.

336

337

338

339 Figure 1b. Density scatterplot comparing CMAQ base daily peak 8-hour ozone predictions against 

340 observed 8-hour ozone peaks paired in space and time for all sites during April-October 2007. 
2A-11 

341

342

343 Figure 1c. Bias in seasonal mean (April-October) maximum daily 8-hour ozone predictions in the 2007 

344 CMAQ base simulation.

345

346

347 Figure 1d. Relationship between CMAQ estimations of MDA8 natural background ozone and daily 

348 model biases. 
2A-12 

349 b. 2007 GEOS-Chem/CAMx source apportionment modeling: 

350 The same global modeling described above was used to assign lateral boundary conditions to 

351 the regional-scale (12 km) CAMx v5.0 source apportionment simulations. Wherever possible, the 

352 emissions and meteorological inputs in the CAMx modeling were chosen to mimic the 2007 base CMAQ 

353 simulation described earlier. Figure 1a shows a map of the model domain.

354 As with the CMAQ base case, a limited operational model performance evaluation was also 

355 completed for surface ozone in the 2007 base simulation. For the purposes of this analysis, EPA 

356 assessed the model ability to reproduce measured daily maximum 8-hour (MDA8) ozone values and 

357 seasonal mean MDA8 ozone concentrations for the period April to October 2007. Figure 2a provides a 

358 density scatterplot of the observed and predicted daily 8-hour ozone peaks paired in space and time for 

359 the 2007 CAMx base simulation. As can be seen, the majority of pairs line up along the 1:1 line. Again, 

360 there is a tendency for the model to overestimate site-days with low 8-hour ozone peaks and 

361 underestimate the site-days with higher peak ozone values. Modeled 8-hour ozone peak concentrations 

362 exhibited relatively small bias and error compared to the observations. The average bias in MDA8 ozone 

363 estimates was 3.5 ppb. Figure 2b depicts the spatial bias patterns in MDA8 ozone at all sites that 

364 measured valid ozone data for at least 100 days during the April-October period. CAMx overestimations 

365 are greatest along the Gulf Coast region, along the Atlantic and Pacific coastlines, and within the 

366 southeastern U.S. The majority of underestimated seasonal mean MDA8 occurs in California away from 

367 the coastline. 

368 The apportionment tools in CAMx utilized here to estimate the contribution of background 

369 sources are well-established and have previously been peer-reviewed (UNC, 2009). EPA used the 

370 Anthropogenic Precursor Culpability Assessment (APCA) tool in this analysis. The APCA tool attributes 

371 ozone production to manmade sources whenever ozone is determined to result from a combination of 

372 anthropogenic and biogenic emissions (Environ, 2011). The APCA methodology defines natural ozone as 

373 the production resulting from the interaction of biogenic VOC with biogenic NOx emissions. Eleven

374 separate source categories were tracked in the source apportionment analysis, including five boundary 

375 condition terms and six in-domain sectors:

376 • Boundary condition terms:

377 o Northern edge

378 o Eastern edge

379 o Southern edge

380 o Western edge

381 o Top boundary

382

383 • In-domain sectors:

384 o U.S. anthropogenic emissions

385 o Point sources located within the Gulf of Mexico

386 o Category 3 marine vessels outside State boundaries

387 o Climatologically-averaged wildfire emissions
2A-13 

388 o Biogenic emissions

389 o Canada/Mexico emissions (only those sources within the domain)

390 It should be noted that the source apportionment modeling conducted here does not allow for 

391 replication of natural background because of the construct of boundary conditions. The boundary 

392 conditions for our applications can include ozone and/or ozone precursors that were originally 

393 generated by natural sources, as well as ozone produced from far upstream anthropogenic emissions 

394 (e.g., Asia). It is not possible to disentangle these two terms. Instead, the source apportionment 

395 modeling is primarily used to help estimate background into the U.S., which is assumed to be the 

396 contributions from nine of the modeled sectors; that is, everything except U.S. anthropogenic emissions 

397 and point sources located within the Gulf of Mexico.

398

399

400 Figure 2a. Density scatterplot comparing CAMx base daily peak 8-hour ozone predictions against 

401 observed 8-hour ozone peaks paired in space and time for all sites during April-October 2007. 

402
2A-14 

403

404 Figure 2b. Bias in seasonal mean (April-October) maximum daily 8-hour ozone predictions in the 2007 

405 CAMx base simulation.

406 3. Estimates of seasonal-average background ozone levels

407 This section of the appendix provides estimates of seasonal average background ozone levels 

408 over the U.S. As noted in the introduction and as discussed in detail in the ISA, background ozone values 

409 can vary significantly in space and time. There can be atypical episodes of higher background ozone 

410 concentrations amidst the routine days that drive seasonal average background. The highest 

411 background episodic concentrations are typically associated with stratospheric intrusions or wildfires. 

412 These background “events” can be difficult to model as they require event-specific model inputs. The 

413 primary goal of the EPA modeling was to estimate the seasonal average background concentrations 

414 between April and October 2007. Previous analyses have shown that this is the period in which average 

415 background levels are highest (Zhang et al., 2011). This section of the appendix focuses on seasonal 

416 mean levels of background. (Section 4 will consider the upper range of possible background ozone.) 

417 The analysis focus on the maximum daily 8-hour ozone average in ppb. This metric is referred to 

418 as MDA8. This section will first present model estimates of seasonal mean ozone levels in the base 

419 simulation. This will be followed by estimates of NB, NAB, and USB from the CMAQ zero out modeling. 

420 After discussing the magnitudes of background levels, the section switches to a consideration of the 

421 relative percentage of background to total ozone across the U.S. This portion of the text will utilize both 

422 the CMAQ zero out and CAMx source apportionment modeling.

423 Figure 3a displays the 2007 base case, CMAQ model-predicted, seasonal mean (April-October) 

424 MDA8 ozone concentrations in grid cells with active monitoring locations over the U.S. The model 

425 results are shown at the monitoring site level as opposed to in the default gridded format to foster 

426 subsequent site-level estimates of background magnitudes. Each grid cell containing an Air Quality 

427 System (AQS) ozone monitor that was collecting valid data in 2007 was identified and the model 
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428 background estimates were extracted for those grid cells and displayed accordingly. The base 

429 predictions are provided for context to allow easier interpretation of the following plots which isolate 

430 specific background levels. As can be seen, most of the U.S. experiences seasonal mean MDA8 ozone 

431 levels greater than 50 ppb in the base case simulation. The median value over the 1,294 monitoring 

432 locations is 52.5 ppb.

433 Figure 3b provides an estimate of what seasonal-average MDA8 would be in a natural 

434 background scenario, using the 2007 EPA zero out modeling. Again, in this GEOS-Chem/CMAQ 

435 simulation, all anthropogenic ozone precursor emissions were removed from both the global and 

436 regional simulations, and methane levels were adjusted to pre-industrial levels in the global simulation. 

437 As shown, natural background ozone levels range from approximately 15-35 ppb with the highest values 

438 occurring over the higher-elevation sites in the western U.S. The median value over these locations is 

439 24.2 ppb, and more than 50 percent of the sites have natural background levels of 20-25 ppb. The 

440 highest modeled estimate of seasonal average, natural background, MDA8 ozone is 34.3 ppb at the 

441 high-elevation CASTNET site (Gothic) in Gunnison County, CO.

442 Figures 3c and 3d show the same information for the North American and U.S. background 

443 scenarios. In these model runs, all anthropogenic ozone precursor emissions were removed from the 

444 U.S., Canada, and Mexico (NAB scenario) and then only the U.S. (USB scenario). The figures show that

445 there is not a large difference between the NAB and USB scenarios. Seasonal mean MDA8 NAB and USB 

446 ozone levels range from 25-50 ppb, with the most frequent values estimated in the 30-35 ppb bin. The 

447 median seasonal mean background levels are 31.5 and 32.7 ppb (NAB and USB, respectively). Again, the 

448 highest levels of background are predicted over the intermountain western U.S. Locations with NAB and 

449 USB concentrations greater than 40 ppb are confined to Colorado, Nevada, Utah, Wyoming, northern 

450 Arizona, eastern California, and parts of New Mexico. Similar to NB, the highest NAB and USB levels 

451 were modeled to occur at the Gothic CO site (46.7/47.7). This remote rural site is located 2,926 meters 

452 (9,600 feet) above mean sea level and should not be considered representative of background ozone at 

453 lower-altitude, more-populated regions. The high USB and NAB values along the Gulf Coast are most 

454 likely due to model biases.

455 Absolute model estimates of various background definitions are useful, but they can be 

456 influenced by any local biases and errors in the modeling. A separate way to look at the role of 

457 background in seasonal mean ozone levels is to consider the fractional contribution of NB, NAB, and USB 

458 to total ozone at each location. Considering the proportional role of background allows for an 

459 informative comparison between the two modeling approaches without having to account for the 

460 differences in base case biases and errors. Figures 4a, 4b, and 4c show the estimated fractional 

461 contribution of NB, NAB, and USB to total seasonal average MDA8 ozone levels at the monitoring 

462 locations from the CMAQ zero out modeling. The modeling estimates that approximately 35-80 percent

463 of the seasonal average MDA8 ozone at monitoring locations is due to natural background sources. A 

464 majority sites have NB fractions between 40 and 60 percent. The mean natural background proportion 

465 over all sites is 47 percent. That is, when all global anthropogenic emissions are removed and global 

466 methane levels in GEOS-Chem are restored to pre-industrial levels, seasonal average MDA8 levels are 

467 reduced by approximately half. The fractional proportions of NAB and USB are very similar. In both 
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468 cases, most sites have background fractions that range from 50 to 80 percent. The mean NAB fraction 

469 (to seasonal mean MDA8) is 63 percent. The mean USB fraction is 66 percent.

470 As noted in the introduction, the advantage of the source apportionment modeling is that all of 

471 the modeled ozone is attributed to various source terms and thus this approach is not affected by the 

472 confounding occurrences of background ozone values exceeding the base ozone values as can happen in 

473 the zero out modeling (i.e., background proportions > 100%). Consequently, one would expect the 

474 fractional background levels to be lower in the source apportionment methodology as a result of 

475 removing this artifact. It is also important to remember that the terms NB, NAB, and USB are explicitly 

476 linked to the zero out modeling approach. (USB is the ozone that would exist in the absence of U.S. 

477 anthropogenic emissions.) In contrast, the source apportionment modeling performed here provides 

478 estimates the amount of MDA8 ozone that is attributable to U.S. anthropogenic emissions relative to 

479 total base model ozone. Figure 4d shows the relative contribution from sources other than U.S. 

480 anthropogenic emissions to total seasonal mean MDA8 ozone based on the 2007 source apportionment 

481 modeling. The fractional contribution fields between CMAQ zero out USB estimates and CAMx source 

482 apportionment estimates of source other than U.S. anthropogenic emissions are quite similar. The 

483 spatial patterns in Figures 4c and 4d are consistent, with the highest fractional contributions from 

484 sources other than U.S. anthropogenic emissions occurring along U.S. borders and over the 

485 intermountain western States. The source apportionment modeling estimates that approximately 40-

486 80% of the seasonal average MDA8 ozone at monitoring locations is due to sources other than 

487 manmade ozone precursor emissions from the U.S. A majority of sites have non-U.S. fractions between 

488 40 and 70 percent. The mean proportion attributable to international and natural sources over all sites 

489 is 59 percent. Despite the differences in the methodologies this is very similar to the mean USB 

490 estimate of 66 percent from the zero out modeling.

491

492

493 
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494

495 Figure 3a. April-October average MDA8 ozone (ppb) at monitoring locations across the U.S. as 

496 estimated by a 2007 CMAQ base simulation.

497

498

499

500 Figure 3b. April-October average natural background MDA8 ozone (ppb) at monitoring locations 

501 across the U.S. as estimated by a 2007 CMAQ zero out simulation.
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502

503 Figure 3c. April-October average North American background MDA8 ozone (ppb) at monitoring 

504 locations across the U.S. as estimated by a 2007 CMAQ zero out simulation.

505

506

507

508 Figure 3d. April-October average United States background MDA8 ozone (ppb) at monitoring 

509 locations across the U.S. as estimated by a 2007 CMAQ zero out simulation.
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510

511 Figure 4a. Ratio of natural background to total April-October average MDA8 ozone at monitoring 

512 locations across the U.S. as estimated based on 2007 CMAQ simulations. 

513

514

515

516 Figure 4b. Ratio of N. American background to total April-October average MDA8 ozone at 

517 monitoring locations across the U.S. as estimated based on 2007 CMAQ simulations.
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518

519 Figure 4c. Ratio of U.S. background to total April-October average MDA8 ozone at monitoring 

520 locations across the U.S. as estimated based on 2007 CMAQ simulations.

521

522

523 Figure 4d. Ratio of sources other than U.S. anthropogenic emissions to total April-October average 

524 MDA8 ozone at monitoring locations across the U.S. as estimated by a 2007 CAMx source 

525 apportionment simulation.
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526 4. Distributions of background ozone levels

527 As a first-order understanding, it is valuable to be able to characterize seasonal mean levels of 

528 background ozone. However, it is well established that background levels can vary substantially from 

529 day-to-day. From an implementation perspective, the values of background ozone on possible 

530 exceedance days is a more meaningful distinction. The first draft policy assessment (EPA, 2012) 

531 considered this issue in detail, via summaries of the existing 2006 zero out modeling (Henderson et al., 

532 2012), and concluded that “results suggest that background concentrations on the days with the highest 

533 total ozone concentrations are not dramatically higher than typical seasonal average background

concentrations.” Based on this finding, the 1st 534 draft policy assessment determined that “anthropogenic 

535 sources within the U.S. are largely responsible for 4th highest 8-hour daily maximum ozone

536 concentrations.” This portion of the appendix will consider the entire spectrum of variable background 

537 ozone levels with special emphasis on days in which base model ozone concentrations approach or 

538 exceed the level of the NAAQS. 

539 The 2007 modeling agrees with the finding from the previous 2006-based modeling analyses 

540 that the highest modeled ozone site-days tend to have background ozone levels similar to mid-range 

541 ozone days. Figures 5a-5c show the distribution of April-October MDA8 background levels (NB, NAB, 

542 and USB, respectively) from the CMAQ zero out runs. As noted in section 2, zeroing out emissions can 

543 remove the effects of local NOx titration and result in modeled background values that are higher than 

544 the base model ozone. The “box and whisker” plots shown in these figures display four key features of 

545 the distributions: 

546 a. the median concentration (black horizontal line) per bin, 

547

b. the inter-quartile range (blue colored box) which represents the 25th-75th 548 percentile range in 

549 values within the distribution, 

550

551 c. the “whiskers” (dark gray vertical lines with top and bottom whiskers) which represent the 

552 range of values within 1.5 times the inter-quartile range, and

553

554 d. the “outliers” (gray points) which are any values outside the whiskers. 

555 As can be seen in Figure 5a, natural background values do not vary greatly as a function of the 

556 base modeled ozone. Recall that the seasonal average natural background MDA8 ozone values were 

557 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 558

559 percentile values are on the order of 30 ppb. Natural background levels exceeding 40-45 ppb are 

560 considered to be statistical outliers, due to their infrequency. Figure 5b shows the same type of 

561 distributions but for NAB instead of NB. NAB values are generally 6-12 ppb higher than their NB 

562 counterparts, due to the affect of higher global methane values and the influence of anthropogenic 

563 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 564 percentile values are 
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565 approximately 40 ppb; it is rare for NAB MDA8 values to exceed 50-55 ppb. NAB values are constant in

566 magnitude once the base ozone exceeds 50 ppb indicating that the higher base ozone values are driven 

567 by non-NAB sources (i.e., North American emissions). Finally in Figure 5c, the USB MDA8 distributions

568 by base model MDA8 are shown. The results are similar to NAB. 

569 Figure 5d shows the results from the source apportionment modeling of non-U.S. anthropogenic 

570 source contributions to MDA8 ozone (i.e., the nine source apportionment categories other than U.S.

571 anthropogenic emissions and Gulf of Mexico point sources). This non-counterfactual approach is 

572 expected to give a better indication of background levels at low concentrations. At low levels, almost all 

573 of the ozone is determined to be from background origins. The CAMx modeling shows that 

574 contributions from non-U.S. anthropogenic emissions peak when base ozone ranges from 45-55 ppb and 

575 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 576

577 percentile values are generally less than 35 ppb, compared to 40 ppb in the zero out modeling. It is rare 

578 to have background impacts greater than 55ppb. Interestingly, when base model MDA8 ozone exceeds 

579 70 ppb, it is rare to have background impacts greater than 45 ppb in the CAMx source apportionment 

580 modeling. 

581 Figures 6a-6d show the equivalent plots as 5a-5d, but use background fractions (background 

582 MDA8 / base MDA8) as the dependent variable instead of the absolute background concentrations. 

583 These plots show the same effect; that is, the proportional relative contribution of background sources 

584 and processes decreases as peak ozone increases. For natural background (Figure 6a), the median 

585 fractions drop from 50% background for values between 45-50 ppb to only 35% background for base 

586 MDA8 values between 70-75 ppb. For NAB and USB (Figures 6b and 6c), the median fractions drop from 

587 70% background for values between 45-50 ppb to only 45% background for base MDA8 values between 

588 70-75 ppb. The source apportionment modeling (Figure 6d) estimates less of a proportional role of non589 U.S. anthropogenic emissions. In that modeling, the median fractions drop from 65% background for 

590 values between 45-50 ppb to only 35% background for base MDA8 values between 70-75 ppb. A key 

591 observation, as noted in the first draft policy assessment document, is that the relative importance of 

592 background decreases on days most likely to violate the NAAQS. An additional policy-relevant finding 

593 from the distributional analyses is that the relative role of background sources would be increased if the 

594 level of the NAAQS were lowered. At 60 ppb, the modeling suggests that the median fractional 

595 contribution from background is 45-55 percent, but there can be cases where background comprises 80-

596 90 percent of the total ozone. 

597 Many of the cases when background ozone is estimated to contribute in large proportions to 

598 relatively high ozone days may be eligible for consideration as exceptional events, but again, this 

599 modeling is not designed to resolve specific events that occurred in 2007. While there is greater 

600 confidence in the model’s ability to predict mean contributions from background sources than from 

601 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 602

603 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 604 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 605 percentile days) are 

observed over the intermountain western U.S. At these locations, 95th 606 percentile USB levels can exceed 

50 ppb. Background values at the 95th 607 percentile end of the distribution are 4-12 ppb higher than the 

608 mean background values at the same locations. 

609

 

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.
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610

611 

612 Figure 5a. Distribution of natural background MDA8 ozone (ppb) at monitoring locations across the 

613 U.S. (Apr-Oct), binned by base modeled site-day MDA8, as estimated by 2007 CMAQ simulations.

614

615 Figure 5b. Distribution of N. American background MDA8 ozone (ppb) at monitoring locations across 

616 the U.S. (Apr-Oct), binned by base modeled site-day MDA8, as estimated by 2007 CMAQ simulations.
2A-25 

617

618 Figure 5c. Distribution of U.S. background MDA8 ozone (ppb) at monitoring locations across the U.S. 

619 (Apr-Oct), binned by base modeled site-day MDA8, as estimated by 2007 CMAQ simulations.

620

621

622 Figure 5d. Distribution of MDA8 ozone contributions from non-U.S. manmade sources (ppb) at 

623 monitoring locations across the U.S. (Apr-Oct), binned by base modeled site-day MDA8, as estimated 

624 by 2007 CAMx simulations.
2A-26 

625

626 Figure 6a. Distribution of natural background MDA8 ozone fractions at monitoring locations across 

627 the U.S. (Apr-Oct), binned by base modeled site-day MDA8, as estimated by 2007 CMAQ simulations.

628

629

630 Figure 6b. Distribution of N. American background MDA8 ozone fractions at monitoring locations 

631 across the U.S. (Apr-Oct), binned by base modeled site-day MDA8, as estimated by 2007 CMAQ 

632 simulations.

633
2A-27 

634

635 Figure 6c. Distribution of U.S. background MDA8 ozone fractions at monitoring locations across the 

636 U.S. (Apr-Oct), binned by base modeled site-day MDA8, as estimated by 2007 CMAQ simulations.

637

638 Figure 6d. Distribution of MDA8 ozone fractions from non-U.S. anthropogenic sources at monitoring 

639 locations across the U.S. (Apr-Oct), binned by base modeled site-day MDA8, as estimated by the 2007 

640 CAMx simulation.

641
2A-28 

642

Figure 7. April-October 95th 643 percentile United States background MDA8 ozone (ppb) at monitoring 

644 locations across the U.S. as estimated by a 2007 CMAQ base simulation.

645

646
2A-29 

647 5. Contribution of various processes and sources to total background ozone 

648 This section will utilize the supplemental 2007 air quality modeling estimates to determine the 

649 relative importance of specific elements of background ozone. Comparing the differences between the 

650 three zero out scenarios can provide some information about the role of certain sets of emissions. 

651 Figure 8a compares the NAB (zero out North American manmade emissions) and USB (zero out U.S. 

652 manmade emissions) scenarios. The difference between these two runs is the inclusion of 

653 anthropogenic emissions within the Canada and Mexico portions of the modeling domain. These 

654 emissions contribute less than 2 ppb to the seasonal mean MDA8 ozone levels over most of the U.S. 

655 There are 70 sites, near an international border, where the modeling estimates Canadian/Mexican 

656 seasonal average impacts of 2-4 ppb. While not shown, the modeled peak single day impacts from 

657 these specific international emissions sources can approach 25 ppb (e.g., San Diego, Buffalo NY). Figure 

658 8b compares the NB (zero out all manmade emissions and reset GEOS-Chem methane values to pre659 industrial levels) to the NAB. The difference between these two runs is the inclusion of global methane 

660 emissions related to recent human activity as well as anthropogenic emissions outside of North America. 

661 These emissions are estimated to contribute 6-15 ppb to seasonal mean ozone levels over the U.S. The 

662 most frequent bin is the 8-10 ppb increase. It is not possible via these runs to parse out what fraction of 

663 this change is due to international emissions as opposed to methane emissions, but the ISA summarized 

664 existing modeling (Zhang et al., 2011) that suggested that the rise in methane from pre-industrial levels 

665 to present-day levels led to increases in seasonal average ozone levels of 4-5 ppb. The greatest impacts 

666 from these sources occurs over the western U.S., where international emissions would be expected to 

667 have the largest impacts.

668 Figures 9a-9g show the fractional contribution to total seasonal mean MDA8 values of 

669 individual source sectors that were tracked in the CAMx source apportionment modeling. Figure 9a 

670 shows the impact from the regional model boundary conditions. The ozone entering the model domain 

671 via the boundary conditions could have a variety of origins including: a) natural sources of ozone and 

672 ozone precursors (including methane) emanating from outside the domain, b) anthropogenic sources of 

673 ozone precursors (including methane) from international emitters, and c) some fraction of U.S. 

674 emissions (natural and anthropogenic) which are exported and then re-imported into the domain via 

675 synoptic-scale recirculation. Thus, one should not presume that the boundary condition contribution is 

676 directly tied to any particular background definition. At most locations, boundary conditions 

677 contributed 40-60 percent of the total MDA8 seasonal mean at sites across the U.S. The highest 

678 proportional impacts from the boundary conditions (the top boundary contributes negligibly) are along 

679 the coastlines and the intermountain West.

680 Figure 9b shows the source apportionment contribution (to seasonal mean MDA8) from the 

681 most significant sector that was tracked: U.S. anthropogenic ozone precursor emissions. Again the most 

682 common outcome at an individual site was that 40-60% of the seasonal mean ozone values originated 

683 from U.S. anthropogenic emissions. The locations with smaller fractional contributions (e.g., 10-20 

684 percent) from U.S. sources are generally located in places where ozone values are typically low such as 

685 the Pacific Northwest. Figures 9c-9g display the fractional contributions from the other five in-domain 

686 sectors listed in section 2. The impacts from these sectors are briefly summarized below:
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687 • Biogenic emissions:

688 o Most frequent bin: 3-5 percent

689 o Highest site-specific contribution: 10-20 percent

690 o Region with greatest impacts: Great Plains states where soil NOx emissions are large 

691 • Climatologically-average fire emissions:

692 o Most frequent bin: 0-1 percent

693 o Highest site-specific contribution: 3-5 percent

694 o Region with greatest impacts: California, Kansas/Oklahoma region

695 • Within-domain Canadian/Mexican manmade emissions:

696 o Most frequent bin: 0-1 percent

697 o Highest site-specific contribution: 10-20 percent

698 o Region with greatest impacts: Sites along international borders (NY, VT, CA, AZ, TX) 

699 • Category 3 marine vessels outside U.S. territorial waters:

700 o Most frequent bin: 0-1 percent

701 o Highest site-specific contribution: 10-20 percent

702 o Region with greatest impacts: Coastal sites (especially southern CA)

• Gulf of Mexico point sources4 703 : 

704 o Most frequent bin: 0-1 percent

705 o Highest site-specific contribution: 1-3 percent

706 o Region with greatest impacts: Sites in southeast TX and southern LA

707

708

709

710

 

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 

711

712 Figure 8a. Difference in April-October average MDA8 ozone (ppb) at monitoring locations across the 

713 U.S. between the USB scenario and the NAB scenario. The difference between these two runs isolates 

714 the impact of within-the-domain anthropogenic emissions from Canada and Mexico. 

715

716 Figure 8b. Difference in April-October average MDA8 ozone (ppb) at monitoring locations across the 

717 U.S. between the NAB scenario and the NB scenario. The difference between these two runs isolates 

718 the impact of the rise in global methane emissions from the pre-industrial and anthropogenic 

719 emissions from outside North America. 
2A-32 

720

721 Figure 9a. Percentage of April-October average MDA8 ozone that is apportioned to boundary 

722 conditions as estimated at monitoring locations by a 2007 CAMx simulation.

723

724

725

726 Figure 9b. Percentage of April-October average MDA8 ozone that is apportioned to U.S. 

727 anthropogenic sources as estimated at monitoring locations by a 2007 CAMx simulation. 
2A-33 

728

729

730 Figure 9c. Percentage of April-October average MDA8 ozone that is apportioned to purely biogenic 

731 emissions as estimated at monitoring locations by a 2007 CAMx simulation.

732

733

734 Figure 9d. Percentage of April-October average MDA8 ozone that is apportioned to climatological fire 

735 emissions as estimated at monitoring locations by a 2007 CAMx simulation. 
2A-34 

736

737 Figure 9e. Percentage of April-October average MDA8 ozone that is apportioned to anthropogenic 

738 emissions from in-domain Canadian and Mexican sources as estimated at monitoring locations by a 

739 2007 CAMx simulation. 

740

741

742 Figure 9f. Percentage of April-October average MDA8 ozone that is apportioned to Category 3 marine 

743 vessel emissions beyond U.S. territorial waters as estimated at monitoring locations by a 2007 CAMx 

744 simulation. 

745
2A-35 

746

747 Figure 9g. Percentage of April-October average MDA8 ozone that is apportioned to Gulf of Mexico 

748 point sources as estimated at monitoring locations by a 2007 CAMx simulation. 

749
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750 6. Estimates of the fractional background contribution to total ozone in 12 specific areas

751 This penultimate section of the appendix presents estimates of the overall fraction of ozone that 

752 is estimated to result from background sources or processes based on the updated modeling in each of 

753 the 12 urban case study areas in the epidemiology study based analyses in Chapter 7 of the Risk and 

754 Exposure Assessment (REA). Tables 1a-1c summarize the CAMx-estimated fractional contributions of 

755 sources other than U.S. anthropogenic emissions to total ozone in each of the 12 areas. Table 1a shows 

756 that the fractional contributions from sources other than anthropogenic emissions within the U.S. to 

757 seasonal mean MDA8 levels can range from 43 to 66 percent across these 12 urban areas. These 

758 fractions are consistent with the national ratios summarized in section 3, although the urban fractions of 

759 background tend to be smaller than at rural sites. As shown in section 4, the fractional contributions 

760 from background are smaller on days with high modeled ozone (i.e., days that may exceed the level of 

761 the NAAQS). Table 1b provides the fractional contributions from these non-U.S. sources, only 

762 considering days in which base model MDA8 ozone was greater than 60 ppb. As expected, the fractional 

763 background contributions are less and range from 31 to 55 percent. Rather than taking the fractions of 

764 the seasonal means (as in Table 1a), Table 1c displays the mean and median daily MDA8 background 

765 fractions. These metrics may be more appropriate for application to health studies, but as can be seen 

766 the fractional contribution to backgrounds calculated via this approach are very similar to the Table 1a 

767 calculations. For completeness sake, although EPA expects the source apportionment results to provide 

768 a more realistic estimate of fractional background values, for completeness, we also provide USB 

769 fractions based on zero out modeling for the 12 cities (see Table 1d). The results are similar to the 

770 source apportionment findings (compare against Table 1a), but the zero out technique provides slightly 

771 higher background proportions.

772

773

774 Table 1a. April-October average MDA8 ozone, average MDA8 ozone from sources other than U.S. 

775 manmade emissions, and the fractional contribution of these background sources in the 12 REA urban 

776 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 

777

778 Table 1b. Average MDA8 ozone, average MDA8 ozone from sources other than U.S. manmade 

779 emissions, and the fractional contribution of these background sources in the 12 REA areas, as 

780 estimated by a 2007 CAMx simulation using site-days in which base MDA8 ozone exceeded 60 ppb.

781

782

783

784 Table 1c. Fractional contribution of non-U.S. manmade emissions sources in the 12 REA urban study 

785 areas, as estimated by a 2007 CAMx simulation using means and medians of daily MDA8 fractions. 

786

787

788

789 Table 1d. April-October average MDA8 ozone, average MDA8 ozone from USB, and the fractional 

790 contribution of these background sources in the 12 REA urban study areas, as estimated by two 

791 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 

792 7. Background ozone and W126

793 As discussed in section 5 of the second draft policy assessment, EPA is considering the adequacy 

794 of the current secondary standard to protect against welfare effects. One metric that has been 

795 considered previously as a potential cumulative seasonal index is the W126 metric. The W126 index is a 

796 sigmoidally weighted sum of all hourly O3 concentrations observed during a specified daily and seasonal 

797 time window, where each hourly O3 concentration is given a weight that increases from 0 to 1 with 

798 increasing concentration (Lefohn et al, 1988). The weights are defined such that values of 0.060 ppm

799 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. 

800 The remainder of this section uses the 2007 zero out modeling to conduct a limited assessment of the 

801 role of background ozone on W126 levels over the U.S.

802 The analysis of background influence on W126 is not as detailed as the analyses related to 

803 seasonal mean MDA8 ozone. Instead of considering impacts at every monitoring location, EPA assessed

804 NB, NAB, and USB influences at four sample locations: Atlanta GA, Denver CO, Farmington NM, and 

805 Riverside CA. Each of these four locations had relatively high observed values of W126 in 2010-2012. 

806 Atlanta is an urban area in the Eastern U.S. with high primary ozone design values but relatively low 

807 levels of seasonal background ozone. Riverside and Denver also have high primary ozone design values 

808 but are in the Western U.S. where background ozone levels are generally higher. Farmington NM was 

809 chosen as a site that has relatively lower primary ozone design values along with its relatively high W126 

810 levels. The varying characteristics of each of these locations perhaps allows broader national 

811 extrapolation of the 4-site results.

812 In previous EPA reviews of the O3 NAAQS, the influence of background ozone was estimated 

813 according to a counterfactual (i.e., how much ozone would exist in the absence of certain sets of 

814 emissions). In the current review, EPA is supplementing the counterfactual assessment with analyses 

815 that estimate the fraction of the existing ozone that is due to background sources. This has important 

816 ramifications for assessing the influence of background on W126 concentrations, because of the non817 linear weighting function used in the metric which emphasizes high ozone hours (e.g., periods in which 

818 ozone is greater than ~60 ppb). As an example, consider a sample site in the intermountain western 

819 U.S. region with very high modeled estimates of U.S background (e.g., seasonal mean USB of 45 ppb 

820 with some days as high as 65 ppb). Even at this high background location, the calculated annual W126 

821 values in the USB scenario are quite low, on the order of 3 ppm-hrs. Most sites in the domain where 

822 background levels are lower than the location cited above will have even smaller background W126 

823 estimates, on the order of 1 ppm-hrs, which is consistent with values mentioned in past reviews (USEPA, 

824 2007). Using the counterfactual scenarios, background ozone has a relatively small impact on W126 

825 levels across the U.S.

826 However, because of the non-linear weighting function used in the W126 calculation, the sum of 

827 the W126 from the USB scenario and the W126 resulting from US anthropogenic sources will not equal 

828 the total W126. In most cases, the sum of those two components will be substantially less than total 

829 W126. As a result, EPA believes it is more informative to estimate the fractional contribution of 

830 background ozone to W126 levels. The 5-step methodology for assessing the fractional influence of 
2A-39 

831 background ozone to annual W126 levels in the four locations is described below. The fractional 

832 influence methodology essentially places higher weights on background fractions on days that are going 

833 to contribute most substantially to the yearly W126 value. 

834 • Step 1a: Calculate the MDA8 ozone values from the base and the three zero out 

835 modeling scenarios at each grid cell containing a site in an area.

836 • Step 1b: Calculate the W126 daily index for the base model scenario.

837 • Step 2: For each site, find the three months with highest summed W126 daily indices.

838 • Step 3: Normalize the daily MDA8 values in the base, NB, NAB, and USB scenarios by the 

839 corresponding W126 daily index from the base scenario. 

840 • Step 4: Calculate the average W126-weighted MDA8 values over the three month 

841 period for each of the four scenarios (base, NB, NAB, USB).

842 • Step 5: Calculate the NB/Base, NAB/Base, and USB/Base ratios based on step 4 outputs. 

843 These values represent an estimate of the fractional influence of background ozone on 

844 modeled W126 levels. 

845 Figure 7a shows the estimated fractional influence of the three background definitions on W126 

846 levels in Atlanta, Denver, Farmington, and Riverside. Based on this limited assessment, natural 

847 background sources are estimated to contribute 29-50% of the total modeled W126 with the highest 

848 relative influence in the intermountain western U.S. (e.g., Farmington NM) and the lowest relative 

849 influence in the eastern U.S. (e.g., Atlanta). U.S. background is estimated to contribute 37-65% of the 

850 total modeled W126. Figure 7b compares the relative influence of background on W126 versus seasonal 

851 mean MDA8 ozone. The proportional impacts of background are slightly less for the W126 metric than 

852 for seasonal mean MDA8 (discussed in section 2.4.2), because of the weighting function that places 

853 more emphasis on higher ozone days when background fractions are generally lower.

854 There are several caveats associated with this analysis. First, only the zero out modeling was 

855 used to assess the fractional influence of background sources on W126. The source apportionment 

856 approach estimated slightly smaller relative contributions for seasonal mean MDA8 levels, so from that 

857 perspective the zero out estimates could represent the high end of background influence on W126. 

858 Additionally, the methodology used for this analysis relies on daily MDA8 values as a surrogate (the data 

859 were readily available) for the 8a-8p time period relevant to the W126 metric. The key conclusion from 

860 this cursory analysis is that background ozone may comprise a non-negligible portion of current W126 

861 levels across the U.S. This fractional influence is greatest in the intermountain western U.S. and are 

862 slightly smaller than the seasonal mean MDA8 metric. In the counterfactual cases, when non

863 background sources are completely removed, the remaining W126 levels are low (< 3 ppm-hrs).
2A-40 

864

865 Figure 7a. Fractional contribution of background sources to W126 levels in four sample locations. 

866 Model estimates based on 2007 CMAQ zero out modeling.

867

868 Figure 7b. Fractional contribution of U.S. background to seasonal mean MDA8 ozone and W126 levels 

869 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 

870 8. Summary 

871 The precise definition of background ozone can vary depending upon context, but it generally 

872 refers to ozone that is formed by sources or processes that cannot be influenced by local control 

873 measures. Background ozone can originate from natural sources of ozone and ozone precursors, as well 

874 as from upwind manmade emissions of ozone precursors. In order to help further characterize 

875 background ozone levels over the U.S., EPA has completed additional air quality modeling analyses 

subsequent to the 1st 876 -draft policy assessment. As shown above, the results are largely consistent with

877 previous determinations about the magnitude of background ozone contributions across the U.S.

878 For a variety of reasons, it is challenging to present a comprehensive summary of all the 

879 components and implications of background ozone. In many forums the term “background” is used 

880 generically and the lack of specificity can lead to confusion as to what sources are being considered. 

881 Additionally, it is well established that the impacts of background sources can vary greatly over space 

882 and time which makes it difficult to present a simple summary of background ozone levels. Further, 

883 background ozone can be generated by a variety of processes, each of which can lead to differential 

884 patterns in space and time, and which often have different regulatory ramifications. Finally, background 

885 ozone is difficult to measure and thus, typically requires air quality modeling which has inherent 

886 uncertainties and potential errors and biases. Even with all of these complexities in mind, EPA believes 

887 the following concise and step-wise summary of background ozone is appropriate as based on previous 

888 modeling exercises and the more recent EPA analyses summarized herein.

889 • The most fundamental definition of background is “natural background” (NB). NB ozone is that 

890 which is produced by processes other than manmade emissions. Examples of sources of natural 

891 background include: stratospheric ozone intrusions, wildfire emissions, and biogenic emissions 

892 from vegetation and soils. To date, NB ozone has been estimated to be that ozone that would 

893 exist in the absence of anthropogenic ozone precursor emissions worldwide. Modeling analyses 

894 have shown that NB levels can vary in time and space. As shown in Section 3, April-October 

895 average NB levels range from approximately 15-35 ppb with the highest values in the spring and 

896 at higher-elevation sites.

897

898 • More expansive definitions of background include North American background (NAB) and U.S. 

899 background (USB). These definitions represent the ozone that originates from sources and 

900 processes other than North American or U.S. anthropogenic sources. Sources of NAB and USB 

901 include all the same sources of natural background, plus manmade ozone precursors emitted 

902 outside the North America or the U.S. Modeling analyses have shown that NAB and USB 

903 background levels can vary in time and space. As discussed in Section 3, seasonal mean NAB 

904 and USB background levels range from approximately 25-45 ppb with the highest values in the 

905 spring and at higher-elevation sites. USB levels are slightly higher than NAB, usually by less than 

906 2 ppb.

907
2A-42 

908 • Estimates of seasonal mean background ozone levels are valuable in terms of a first-order 

909 characterization, however because levels can vary significantly from day-to-day, it is also 

910 instructive to consider the distribution of daily model estimates of background ozone over a 

911 season. Typically, model background is slightly higher in the April-June period than in the later 

912 portion of the ozone season (July-October) (EPA, 2012). More importantly, the modeling shows 

913 that the days with highest ozone levels, on average, have similar background levels to days with 

914 lower values. As a result, the proportion of total ozone that has background origins is smaller 

915 on high ozone days (e.g., days > 70 ppb) than the more common lower ozone days that drive 

916 seasonal means. Section 4 provides information about the distribution of background ozone 

917 fractions. Based on the source apportionment modeling, it is shown that U.S. anthropogenic 

918 emissions typically comprise the majority of the total ozone on site-days with base modeled 

919 ozone MDA8 values greater than 60 ppb.

920

921 • While it is important to recognize that most high ozone days (i.e., potential exceedance days) 

922 are estimated to be driven predominantly by non-background emissions, the recent EPA 

923 modeling also shows times and locations in which background contributions are estimated to 

924 approach 60-80 ppb. As described in Sections 4 and 6 of this document, these occurrences are 

925 relatively infrequent. While the modeling was not expressly developed to capture these types 

926 of events, ambient observations have also shown relatively rare events where background 

927 ozone sources (wildfires, stratospheric intrusions) have overwhelmingly contributed to an 

928 ozone exceedance. From a policy perspective, these background events must be viewed in the 

929 context of their relative infrequency and the existing mechanisms within the Clean Air Act (e.g., 

930 exceptional event policy, 179B international determinations) that help ensure States are not 

931 required to control for events that are inherently outside their ability to influence. While 

932 background ozone levels can approach and periodically exceed the NAAQS at some locations, 

933 these conditions are not a constraining factor in the selection of a NAAQS. The Clean Air Act 

934 requires the NAAQS to be set at a level requisite to protect public health and welfare. Case law 

935 makes it clear that attainability and technical feasibility are not relevant considerations. In 

936 previous reviews, EPA assessed the proximity of potential levels to peak background levels as a 

937 secondary consideration between levels where health and welfare was protected.

938

939 • Section 5 shows that the contributions to background are multi-dimensional. Daily peak 8-hour 

940 ozone values over the U.S. are a function of local and regional anthropogenic emissions, 

941 anthropogenic emissions from outside the U.S. (including shipping emissions), natural and 

942 anthropogenic methane emissions, wildfire emissions, and purely natural sources. While local 

943 and regional controls are still considered to be the most effective at reducing local ozone levels, 

944 any measures to reduce the international contributions or methane-induced background will 

945 also be valuable.

946

947 • In previous ozone NAAQS reviews, EPA estimated risk from exposure only to ozone 

948 concentrations above background. In the first drafts of the REA and PA for the current ozone 
2A-43 

949 review, EPA estimated risk from exposure to total measured ozone concentrations, which 

950 include those concentrations from background sources. EPA will continue to provide estimates 

951 of risk from exposure to total ozone, consistent with CASAC advice, in the second draft policy 

952 assessment. The recent EPA modeling was completed to assist in determining, in a limited 

953 sense, the risk attributable to background ozone. The fractional values of background 

954 contributions in the 12 REA study areas (43-66 percent) could be used as first order 

955 approximations of the risk due to ozone background. 

956

957 9. References 

958 Arunachalam, S. (2009). Peer Review of Source Apportionment Tools in CAMx and CMAQ. Prepared 

959 under EPA Contract #EP-D-07-102, Institute of the Environment, University of North Carolina, 

960 Chapel Hill, NC, 42pp, 

961 http://www.epa.gov/scram001/reports/SourceApportionmentPeerReview.pdf. 

962 Emery, C; Jung, J; Downey, N; Johnson, J; Jimenez, M; Yarwood, G; Morris, R. (2012). Regional and global 

963 modeling estimates of policy relevant background ozone over the United States. Atmospheric 

964 Environment, 47: 206-217. http://dx.doi.org/10.1016/j.atmosenv.2011.11.012. 

965 Environ (2011). User’s Guide: Comprehensive Air Quality Model with Extensions, Version 5.40; Novato 

966 CA, 306pp. http://www.camx.com/files/camxusersguide_v5-40.aspx. 

967 Fiore, A; Jacob, DJ; Liu, H; Yantosca, RM; Fairlie, TD; Li, Q. (2003). Variability in surface ozone background 

968 over the United States: Implications for air quality policy. J Geophys Res, 108: 4787. 

969 http://dx.doi.org/10.1029/2003JD003855. 

970 Henderson, BH; Possiel N; Akhtar F; Simon H. (2012). Memo to the Ozone NAAQS Review Docket EPA971 HQ-OAR-2012-0699: Regional and Seasonal Analysis of North American Background Ozone 

972 Estimates from Two Studies. 

973 http://www.epa.gov/ttn/naaqs/standards/ozone/s_o3_2008_td.html. 

974 Lefohn, A. S.; Laurence, J. A.; Kohut, R. J. (1988). A comparison of indices that describe the relationship 

975 between exposure to ozone and reduction in the yield of agricultural crops. Atmos. Environ. 22: 

976 1229-1240. 

977 Simon, H., Baker, K.P., Phillips, S. (2012) Compilation and interpretation of photochemical model 

978 performance statistics published between 2006 and 2012. Atmospheric Environment, 61, 124-

979 139.

980 U.S. Environmental Protection Agency (2005). Technical Support Document for the Final Clean Air 

981 Interstate Rule Air Quality Modeling. Office of Air Quality Planning and Standards, Research 

982 Triangle Park, NC, 285pp. http://www.epa.gov/cair/technical.html. 

983 U.S. Environmental Protection Agency. (2007). Review of the National Ambient Air Quality Standards for 

984 Ozone: Policy Assessment of Scientific and Technical Information - OAQPS Staff Paper, U.S. 

985 Environmental Protection Agency, Research Triangle Park, NC. EPA 452/R-07-007.
2A-44 

986 U.S. Environmental Protection Agency (2011). Air Quality Modeling Final Rule Technical Support 

987 Document. Office of Air Quality Planning and Standards, Research Triangle Park, NC, 363pp. 

988 http://www.epa.gov/airtransport/CSAPR/techinfo.html. 

989 U.S. Environmental Protection Agency (2012). Policy Assessment for the Review of the Ozone National 

990 Ambient Air Quality Standards: First External Review Draft. EPA-452/P-12-002, 297pp.

991 U.S. Environmental Protection Agency. (2013). Integrated Science Assessment for Ozone and Related 

992 Photochemical Oxidants, U.S. Environmental Protection Agency, Research Triangle Park, NC. 

993 EPA/600/R-10/076.

994 U.S. Environmental Protection Agency. (2014). Health Risk and Exposure Assessment for Ozone, Second 

995 External Review Draft, U.S. Environmental Protection Agency, Research Triangle Park, NC. EPA 

996 xxx/P-xx-xxx.

997 Zhang, L; Jacob, DJ; Downey, NV; Wood, DA; Blewitt, D; Carouge, CC; Van donkelaar, A; Jones, DBA; 

998 Murray, LT; Wang, Y. (2011). Improved estimate of the policy-relevant background ozone in the 

999 United States using the GEOS-Chem global model with 1/2 2/3 horizontal resolution over North 

1000 America. Atmospheric Environment 45: 6769-6776. 

1001 http://dx.doi.org/10.1016/j.atmosenv.2011.07.054. 
2B-1

1 APPENDIX 2B

2 MONITORING DATA ANALYSIS OF RELATIONSHIPS 

3 BETWEEN CURRENT STANDARD AND W126 METRIC

4 Presented here are monitoring data analyses evaluating relationship between ozone (O3) 

5 concentrations in the averaging time and form of the current secondary standard (3-year average 

6 of the annual 4th highest daily maximum 8-hour concentrations, in parts per billion), and a three7 year W126 metric (3-year average of the annual maximum 3-month sum of weighted daytime 

8 concentrations, in parts per million-hours). We also consider the responsiveness of these two 

9 metrics to historical changes in air quality related to ozone precursor emissions. 

10 For this analysis, we chose to examine monitoring data from a base period (2001-2003) 

11 as well as a recent period (2009-2011). The base period was chosen to represent air quality 

12 conditions before the implementation of the 1997 national ambient air quality standard 

13 (NAAQS) for O3 (0.08 ppm). In 2004, EPA designated 113 areas as nonattainment for the 1997 

14 standard, which required many areas to begin precursor emissions control programs for the first 

15 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 16 ,” which required 

17 summertime reductions in NOx emissions from power plants and other large sources throughout 

18 the Eastern U.S. These programs were successful in reducing peak O3 concentrations, especially 

19 in the Eastern U.S., and as a result only 8 of the original 113 nonattainment areas were still

20 violating the 1997 O3 NAAQS during the 2009-2011 period. 

21 Hourly O3 concentration data were retrieved from EPA’s Air Quality System (AQS) 

database2 22 for both periods, and used to calculate design values for the current standard as well as 

23 3-year average W126 values for both periods. The procedures for calculating design values for 

24 the current standard from hourly O3 concentration data are described in 40 CFR Part 50, 

25 Appendix P, and the procedures for calculating the 3-year average W126 values are described in 

26 section 4.3.1. of the 2nd draft Welfare Risk and Exposure Assessment (WREA). There were 838 

27 monitoring sites with sufficient data to calculate these values for both periods. In order to 

28 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

29 climate regions (Karl and Koss, 1984) used in the WREA. Figure 2B-1 presents a map of these 

30 regions, which are color-coded to match the scatter plots in the subsequent figures. 

31 Figures 2B-2a, 2B-2b, 2B-3a and 2B-3b show scatter plots of the design values for the 

32 current standard (x-axis) versus 3-year average W126 values (y-axis) for the base period and 

33 recent period, respectively. Most monitors in the U.S. both exceeded the current standard of 75 

34 ppb and a three-year average W126 value of 15 ppm-hrs during the base period. During the 

35 recent period, both the design values and 3-year average W126 values were much lower, and 

36 there also appears to be less scatter between the two metrics. In both periods, the highest design 

37 values and W126 values occurred in the West region which includes California. Finally, it is 

38 worth noting that monitors in the Southwest and West regions tend to have higher W126 values 

39 relative to their design values than in other regions. 

40 Figure 2B-4 shows a scatter plot of the design values for the current standard for the base 

41 period (x-axis) versus for the recent period (y-axis), while Figure 2B-5 shows this same 

42 relationship based on the 3-year average W126 values. The relationship between the two periods 

43 appears to be fairly linear for both metrics, indicating that larger decreases in these metrics 

44 tended to occur at monitors with higher base values. Figures 2B-6 and 2B-7 show design values 

45 for the current standard and 3-year average W126 values, respectively, compared to the unit 

46 changes in those values between the base period and recent period. Figures 2B-6 and 2B-7 show 

47 the difference between each point and the one-to-one lines in Figures 2B-4 and 2B-5, 

48 respectively. In particular, these figures highlight that there were some monitors where design 

49 values for the current standard and/or W126 values increased. However, those monitors also

50 tended to have lower base values, and were mostly located outside of areas subject to emissions 

51 controls under the 1997 standard. 

52 Finally, Figure 2B-8 compares the unit change in design values (in ppb; x-axis) to the 

53 unit change in 3-year average W126 values (in ppm-hrs; y-axis). This figure shows that in most 

54 locations, the current standard metric and the W126 metric exhibit similar responses to changes 

55 in precursor emissions. In particular, the NOx SIP Call, which was implemented in the states 

56 east of the Mississippi River, was effective at reducing both design values and W126 values at 

57 nearly all monitors in the Eastern U.S. The relationship was much more variable in the 

58 remaining regions, where emissions control programs were mostly local and limited to areas

59 which were violating the NAAQS.

60 Based on this analysis of ambient monitoring data, we can make the following general 

61 conclusions about the relationship between the design value metric for the current O3 standard 

62 and the 3-year average W126 metric: 

63 1. There is a fairly strong, positive degree of correlation between the two metrics.
2B-3

64 2. Monitors in the West and Southwest regions tend to have higher W126 values relative to 

65 their design values than in other regions. 

66 3. Reducing precursor emissions, especially NOx, is an effective strategy for lowering both 

67 design values and W126 values. In particular, regional control programs such as the NOx 

68 SIP call are effective at reducing both metrics over a broad area.

69

70 In addition, Figure 2B-9 examines the number of counties with 8-hour design values 

71 meeting the current standard and 3-year average W126 index values greater than 15 ppm-hrs.

72 Most of these counties were located in the Southwest region of the country. There were no 

73 counties in any of the studied 3-year periods that had design values less than or equal to 65 ppb 

74 and 3-year average W126 index values greater than 15 ppm-hrs.

75

76

77 Figure 2B-1. Map of the 9 NOAA climate regions (Karl and Koss, 1984), color coded to 

78 match the subsequent scatter plots. 
2B-4

79

80 Figure 2B-2a. Design values for the current O3 standard in ppb (x-axis) versus 3-year 

81 average W126 values in ppm-hrs (y-axis) based on ambient monitoring data 

82 for 2001-2003. 

83
2B-5

84

85

86 Figure 2B-2b. Design values for the current O3 standard in ppb (x-axis) versus 3-year 

87 average W126 values in ppm-hrs (y-axis) based on ambient monitoring data 

88 for 2001-2003 with a focus on monitors with 2001-2003 design values below 75 

89 ppb. 

90
2B-6

91

92 Figure 2B-3a. Design values for the current O3 standard in ppb (x-axis) versus 3-year 

93 average W126 values in ppm-hrs (y-axis) based on ambient monitoring data 

94 for 2009-2011. 

95
2B-7

96

97

98 Figure 2B-3b. Design values for the current O3 standard in ppb (x-axis) versus 3-year 

99 average W126 values in ppm-hrs (y-axis) based on ambient monitoring data 

100 for 2009-2011 with a focus on monitors with 2009-2011 design values below 75 

101 ppb.

102
2B-8

103

104 Figure 2B-4. Design values for the current O3 standard in ppb based on ambient 

105 monitoring data for 2001-2003 (x-axis) versus 2009-2011 (y-axis). 
2B-9

106

107 Figure 2B-5. Three-year average W126 values in ppm-hrs based on ambient monitoring 

108 data for 2001-2003 (x-axis) versus 2009-2011 (y-axis). 

109
2B-10

110

111 Figure 2B-6. Design values for the current O3 standard in ppb based on ambient 

112 monitoring data for 2001-2003 (x-axis) versus unit (ppb) change in design 

113 values from 2001-2003 to 2009-2011 (y-axis). 

114
2B-11

115

116 Figure 2B-7. Three-year average W126 values in ppm-hrs based on ambient monitoring 

117 data for 2001-2003 (x-axis) versus unit (ppm-hr) change in 3-year average 

118 W126 values from 2001-2003 to 2009-2011 (y-axis). 

119
2B-12

120

121 Figure 2B-8. Unit (ppb) change in design values for the current O3 standard from 2001-

122 2003 to 2009-2011 (x-axis) versus unit (ppm-hr) change in 3-year average 

123 W126 values from 2001-2003 to 2009-2011 (y-axis). 

124

125
January 2014 2B-13 Draft - Do Not Quote or Cite

126

127

128 Figure 2B-9. Number of counties where the 8-hour design value is meeting the current standard and 3-year average W126 

129 index value is greater than 15 ppm-hrs (left), and number of counties where the 8-hour design value is less than 

or equal to 70 ppb and 3-year average W126 index value is greater than15 ppm-hrs (right)3 130 . 

131

132 REFERENCES

133 Karl, T.R. and Koss, W.J., 1984: "Regional and National Monthly, Seasonal, and Annual Temperature Weighted by Area, 1895-1983." Historical Climatology 

134 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 averages of 65 ppb and also 

above a 3-year W126 index value of 15 ppm-hrs.
2C-1 

1 APPENDIX 2C

2 INTER-ANNUAL VARIABILITY IN W126 INDEX VALUES: 

3 COMPARING ANNUAL AND 3-YEAR AVERAGE METRICS (2008-2010)

4 2C.1 OVERVIEW

5 This appendix describes an analysis comparing values for a single-year or annual W126 

6 metric to a W126 metric averaged over three consecutive years. The purpose of this analysis is 

7 to compare values based on a 3-year average of annual W126 indices to values based on a single 

8 annual W126 index. The deviations of the annual W126 index values in 2008, 2009, and 2010 

9 from the 2008-2010 average W126 index values are presented.

10 2C.2 GENERAL DATA PROCESSING

11 The air quality data for this analysis originated from EPA’s Air Quality System (AQS) 

12 data base, the official repository of ambient air measurements. The data used in this analysis 

13 consisted of W126 index values calculated from hourly ozone concentrations measured at 1082 

14 ozone monitors nationwide. Ozone monitors must have submitted data to AQS for at least 75% 

15 days in their required ozone monitoring season in 2008, 2009, and 2010 to be included in the 

16 analysis.

17 2C.3 RESULTS & CONCLUSION

18 The figure below shows a scatter plot of the deviations in the annual W126 index from 

19 the 3-year average by monitor. The solid curves represent the average deviation in a moving 

20 window along the x-axis for each year. From this figure, it is apparent that the highest annual 

21 W126 index value occurred in 2008 for most monitoring locations, the lowest annual W126 

22 index value occurred in 2009 for most monitoring locations, and the 2010 W126 index value was 

23 generally somewhere in between. It is also apparent that the inter-annual variability in the W126 

24 index increases along with the 3-year average. For monitors with 3-year average W126 values 

25 near 15 ppm-hrs, the average deviation was +3.5 ppm-hrs in 2008 and -3.8 ppm-hrs in 2009. 

26 This represents a 1-year swing of -7.3 ppm-hrs.

The model-based air quality adjustments in the 2nd 27 draft of the O3 Welfare REA show 

28 that reducing NOx emissions is effective for reducing 3-year average W126 levels. In Appendix 

29 2B, the analyses based on ambient monitoring data also show that large-scale reductions in NOx

30 emissions are associated with lower W126 levels. Finally, the data analysis presented in this 

31 appendix shows that the inter-annual variability in the annual W126 index tends to decrease with 

32 decreasing W126 levels. Thus, it is expected that reductions in NOx emissions will not only 
2C-2 

33 result in lower 3-year average W126 levels, but also result in less inter-annual variability 

34 associated with annual W126 levels. 

35 The W126 index is based on a logistic weighting function that increases the weights 

36 assigned to hourly ozone concentrations very rapidly. Hourly ozone concentrations of 50 parts 

37 per billion are given a weight of about 10% while concentrations of 80 parts per billion are given 

38 a weight of nearly 90%. The annual W126 index is calculated as a 3-month sum of weighted 

39 ozone concentrations during daylight hours, which amounts to a sum of roughly 1100 weighted 

40 hourly concentrations. Thus, even a modest change in the average daily ozone level may have a 

41 significant impact upon the annual W126 index. Since ozone formation is heavily influenced by 

42 meteorology, the inter-annual variability in meteorological conditions tends to cause a large 

43 inter-annual variability in the W126 index. 

44 In conclusion, this evaluation indicates the extent to which a form for the secondary 

45 ozone standard that averages the annual W126 index values over three consecutive years might 

46 be expected to account for the annual variability in this index since the 3-year period would be 

47 expected to include year(s) below as well as above the 3-year average.

48 
2C-3 

49 

50 

51 Figure 2C-1. Deviation of the annual W126 index values in 2008, 2009, and 2010 (y-axis) 

52 from the 3-year average W126 index value (x-axis). 
3A-1 

1 APPENDIX 3A

2 MODES OF ACTION SUMMARY

3 The initial key event in the toxicity pathway of O3 is the formation of secondary 

4 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)

1 5 . The resulting 

6 secondary oxidation products transmit signals to the epithelium, pain receptive nerve fibers and, 

7 if present, immune cells (i.e., eosiniphils, dendritic cells and mast cells). Thus, the effects of O3

8 are mediated by components of ELF and by the multiple cell types found in the respiratory tract. 

Further, oxidative stress2 9 is an implicit part of this initial key event.

10 Another key event in the toxicity pathway of O3 is the activation of neural reflexes which 

11 lead to lung function decrements. Evidence is accumulating that secondary oxidation products 

12 are responsible for this effect. Different receptors on bronchial sensory nerves (i.e., C-fibers) 

13 have been shown to mediate separate effects of O3 on pulmonary function. For example, pain 

14 (i.e., nociceptive) sensory nerves are involved in the involuntary truncation of inspiration which 

15 results in decreases in FVC, FEV1, tidal volume and pain upon deep inspiration. Ozone exposure 

16 also results in activation of vagal sensory nerves and a mild increase in airway obstruction 

17 measured as increased sRaw. Activation of neural reflexes also results in extrapulmonary effects 

18 such as slow resting heart rate (i.e., bradycardia). 

19 Initiation of inflammation is also a key event in the toxicity pathway of O3. Secondary 

20 oxidation products, as well as cell signaling molecules (i.e., chemokines and cytokines) from

21 airway epithelial cells and white blood cells (i.e., macrophages), have been implicated in the 

22 initiation of inflammation. Airways neutrophilia has been demonstrated in bronchoalveolar 

23 lavage fluid (BALF), mucosal biopsy and induced sputum samples. Influx of other cells (i.e. 

24 mast cells, monocytes and macrophages) also occur. Inflammation further contributes to 

25 O3-mediated oxidative stress. It should be noted that inflammation, as measured by airways 

 1 The ELF is a complex mixture of lipids (fats), proteins, and antioxidants that serve as the first barrier and 

target for inhaled O3. The 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 ELF.

2 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.
3A-2 

1 neutrophilia, is not correlated with decrements in pulmonary function as measured by 

2 spirometry.

3 A fourth key event in the toxicity pathway of O3 is alteration of epithelial barrier

function. Increased permeability3 4 occurs as a result of damage to tight junctions between 

5 epithelial cells subsequent to O3-induced injury and inflammation. It may play a role in allergic 

6 sensitization and in airway hyperresponsiveness (AHR). Genetic susceptibility has been 

7 associated with this pathway. 

8 A fifth key event in the toxicity pathway of O3 is the sensitization of bronchial smooth 

9 muscle. Airway hyperresponsiveness (AHR), or increased bronchial reactivity, can be both a 

10 rapidly occurring and a persistent response. The mechanisms responsible for early and later AHR 

11 are not well-understood. Tachykinins, peptides that can excite neurons and cause smooth muscle 

12 contraction, and the secondary oxidation products of O3 have been proposed as mediators of the 

13 early response and inflammation-derived products have been proposed as mediators of the later 

14 response. Other chemical signaling molecules (i.e., cytokines and chemokines) have been 

15 implicated in the AHR response to O3 in animal models. Antioxidants may confer protection. 

16 A sixth key event in the toxicity pathway of O3 is the modification of innate/adaptive 

17 immunity. While the majority of evidence for this key event comes from animal studies, there 

18 are several studies suggesting that this pathway may also be relevant in humans. Ozone exposure 

19 of human subjects resulted in recruitment of activated innate immune cells to the airways. 

20 Animal studies further linked O3-mediated activation of the innate immune system to the 

21 development of nonspecific AHR, demonstrated an interaction between allergen and O3 in the 

22 induction of nonspecific AHR, and found that O3 acted as an adjuvant for allergic sensitization 

23 through the activation of both innate and adaptive immunity. These studies provide evidence that 

24 O3 can alter host immunologic response and lead to immune system dysfunction. These 

25 mechanisms may underlie the exacerbation and induction of asthma, as well as decreases in lung 

26 host defense.

27 Another key event in the toxicity pathway of O3 is airways remodeling. Persistent 

28 inflammation and injury, which are observed in animal models of chronic and intermittent 

exposure to O3, are associated with morphologic changes such as mucous cell metaplasia4 29 of 

 3 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.

4 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 
3A-3 

1 nasal epithelium, bronchiolar metaplasia of alveolar ducts and fibrotic changes in small airways 

2 (see Section 7.2.3 of the ISA, U.S. EPA 2013). Mechanisms responsible for these responses are 

3 not well-understood. However, a recent study in mice demonstrated a key role for a signaling 

4 pathway in the deposition of collagen in the airway wall following chronic intermittent exposure 

5 to O3. Chronic intermittent exposure to O3 has also been shown to result in effects on the 

6 developing lung and immune system. 

7 Systemic inflammation and vascular oxidative/nitrosative stress are also key events in the 

8 toxicity pathway of O3. Extrapulmonary effects of O3 occur in numerous organ systems, 

9 including the cardiovascular, central nervous, reproductive, and hepatic systems (see 

10 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 

11 lipid oxidation products resulting from reaction of O3 with lipids and/or cellular membranes in 

12 the ELF are responsible for systemic responses; however, it is not known whether they gain 

13 access to the circulation. Alternatively, release of diffusible mediators from the lung into the 

14 circulation may initiate or propagate inflammatory responses in the circulation or other organ 

15 systems. 

16 Responses to O3 exposure are variable within the population. Studies have shown a large 

17 range of pulmonary function responses to O3 among healthy young adults. Since individual 

18 responses were relatively consistent across time in some of these studies, it was thought that 

19 responsiveness reflected an intrinsic characteristic of the subject (Mudway and Kelly, 2000). 

20 Other responses to O3 have also been characterized by a large degree of inter-individual 

21 variability. For example, inter-individual variability in the neutrophilic response has been noted 

22 in human subjects. Two studies demonstrated a 3- to 20-fold difference in airways neutrophilia, 

23 under different exposure conditions (Schelegle et al., 1991 and Devlin et al., 1991, respectively). 

24 Reproducibility of intraindividual responses to O3 exposures in human subjects, measured as 

25 sputum neutrophilia, was demonstrated by Holz et al (1999). While the basis for the observed 

26 inter-individual variability in responsiveness to O3 is not clear, both dosimetric and mechanistic 

27 factors are likely to contribute. These are discussed in sections 5.4.1 and 5.4.2 in the ISA (U.S. 

28 EPA, 2013). Section 5.4.2 of the ISA discusses studies that provide evidence for the mechanisms 

29 that may underlie the variability in responsiveness seen among individuals. Certain functional 

30 genetic polymorphisms, pre-existing conditions or diseases, nutritional status, lifestages, and co31 exposures contribute to altered risk of O3-induced effects.

32 

 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.
3A-4 

1 3.1 REFERENCES

2 Devlin, RB; Mcdonnell, WF; Mann, R; Becker, S; House, DE; Schreinemachers, D; Koren, HS. (1991). Exposure of 

3 humans to ambient levels of ozone for 6.6 hours causes cellular and biochemical changes in the lung. Am J 

4 Respir Cell Mol Biol 4: 72-81.

5 Holz, O; Jorres, RA; Timm, P; Mucke, M; Richter, K; Koschyk, S; Magnussen, H. (1999). Ozone-induced airway 

6 inflammatory changes differ between individuals and are reproducible. Am J Respir Crit Care Med 159: 

7 776-784.

8 Mudway, IS; Kelly, FJ. (2000). Ozone and the lung: A sensitive issue. Mol Aspects Med 21: 1-48.

9 Schelegle, ES; Siefkin, AD; McDonald, RJ. (1991). Time course of ozone-induced neutrophilia in normal humans. 

10 Am J Respir Crit Care Med 143: 1353-1358.

11 U.S. EPA (2013). Integrated Science Assessment of Ozone and Related Photochemical Oxidants (Final Report). 

12 U.S. Environmental Protection Agency, Washington, DC. EPA/600/R-10/076F. Available at: 

13 http://www.epa.gov/ttn/naaqs/standards/ozone/s_o3_2008_isa.html 
3B-1 

1 APPENDIX 3B

2 RECENT STUDIES OF RESPIRATORY-RELATED 

3 EMERGENCY DEPARTMENT VISITS AND HOSPITAL 

4 ADMISSIONS

5 Hospital Admissions for All Respiratory Causes

6 The APHENA study (APHENA is for Air Pollution and Health: A European and North 

7 American Approach) analyzed air pollution and health outcome data from existing Canadian, 

8 European, and U.S. multi-city studies and examined the influence of varying model specification 

9 to control for season and weather (Katsouyanni et al., 2009). The U.S.-based portion of the 

10 APHENA study utilized the National Morbidity, Mortality, and Air Pollution Study (NMMAPS) 

11 cohort which, for the Katsouyanni et al. (2009) analysis, comprised respiratory hospital 

12 admissions among individuals 65 years of age and older from 14 US cities with O3 data from 

13 1985-1994 (7 cities had summer only O3 data). For the year round analysis, Katsouyanni et al. 

14 (2009) reported consistently positive, and statistically significant in models with 8 degrees of 

15 freedom per year (U.S. EPA, 2013, section 6.2.7.2), associations between 1-hour O3

16 concentrations and respiratory hospital admissions across the datasets from the U.S., Canada, and 

Europe (U.S. EPA 2013, Figure 6-15).5 17 In co-pollutant models adjusting for PM10, O3 effect 

18 estimates remained positive, though effect estimates were somewhat attenuated in the U.S. and 

19 European datasets, possibly due to the PM sampling schedule (U.S. EPA 2013, Figure 6-15). 

20 Effect estimates for the warm season were larger than for the year-round analysis in the 

21 Canadian dataset, but generally similar in magnitude to the year-round analysis in the U.S. and 

22 European datasets. 

23 Several additional multicity studies examined respiratory disease hospital admissions in 

24 Canada and Europe. Cakmak et al. (2006) reported a statistically significant increase in 

25 respiratory hospital admissions in 10 Canadian cities (4.4% increase per 20 ppb increase in 24-

26 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 27 

28 grade were found to be at greater risk. Dales et al. (2006) reported a 5.4% (95% CI: 2.9, 8.0%) 

29 increase in neonatal respiratory hospital admissions for a 20 ppb increase in 24-hour average O3

 5

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). 
3B-2 

1 concentrations in 11 Canadian cities from 1986 to 2000. In contrast, Biggeri et al. (2005) did not 

2 detect an association between short-term O3 exposure and respiratory hospital admissions in four 

3 Italian cities from 1990 to 1999. 

4 In addition to the large multi-city studies discussed above, several smaller-scale studies 

5 have also reported associations with total respiratory hospital admissions. Specifically, Lin et al. 

6 (2008) reported a positive association between O3 and pediatric (i.e., <18 years) respiratory 

7 admissions in an analysis of 11 geographic regions in New York state from 1991 to 2001, though 

8 results were not presented quantitatively. In co-pollutant models with PM10, the authors reported 

9 that region-specific O3 associations with respiratory hospital admissions remained relatively 

10 robust. 

11 Cause-Specific Hospital Admissions

12 With regard to cause-specific respiratory outcomes, the limited evidence available in the 

13 last review indicated that the strongest findings were for ambient O3 associated asthma and 

14 chronic obstructive pulmonary disease (COPD) respiratory hospital admissions (U.S. EPA 2013, 

15 6.2.7.2). Since the last review, a few additional studies have investigated cause-specific 

16 respiratory admissions (i.e., COPD, asthma, pneumonia) in relation to O3 exposure (Medina17 Ramon et al, 2006; Yang et al., 2005; Zanobetti and Schwartz, 2006; Silverman and Ito, 2010). 

18 Medina-Ramon et al. (2006) examined the association between short-term ambient O3

19 concentrations and Medicare hospital admissions for COPD among individuals ≥ 65 years of age 

20 for COPD in 35 cities in the U.S. for the years 1986-1999. The authors reported an increase in 

21 COPD admissions for lag 0-1 day in the warm season for a 30 ppb increase in 8-h max O3

22 concentrations. The authors found no evidence for such associations in cool season or in year 

23 round analyses. In a co-pollutant model with PM10, the association between O3 and COPD 

24 hospital admissions remained robust. In Vancouver from 1994-1998, a location with low ambient 

25 O3 concentrations (U.S. EPA, 2013, Table 6-26), Yang et al. (2005) reported a statistically non26 significant increase in COPD admissions per 20 ppb increase in 24-hour average O3

27 concentrations. In two-pollutant models with every-day data for NO2, SO2, CO, and PM10, O3

28 risk estimates remained robust, though not statistically significant (U.S. EPA, 2013, Figure 6-20; 

29 Table 6-29). In addition, Wong et al. (2009) reported increased O3-associated COPD admissions 

30 during periods of increased influenza activity in Hong Kong. 

31 The ISA assessed a study that evaluated asthma-related hospital admissions in New York 

32 City (U.S. EPA, 2013, section 6.2.7.2) (Silverman and Ito, 2010). This study examined the 

33 association of 8-hour max O3 concentrations with severe acute asthma admissions (i.e., those 

34 admitted to the Intensive Care Unit [ICU]) during the warm season in the years 1999 through 
3B-3 

1 2006 (Silverman and Ito, 2010)). The investigators reported positive associations between O3 and 

2 ICU asthma admissions for the 6- to 18-year age group for a 30 ppb increase in max 8-hour 

3 average O3 concentrations, but little evidence of associations for the other age groups examined 

4 (<6 years, 19-49, 50+, and all ages). However, positive associations were observed for each 

5 age-stratified group and all ages for non-ICU asthma admissions, but again the strongest 

6 association was reported for the 6- to 18-years age group. In two-pollutant models, O3 effect 

7 estimates for both non-ICU and ICU hospital admissions remained robust to adjustment for 

8 PM2.5. In an additional analysis, using a smooth function, the authors examined whether the 

9 shape of the concentration-response curve for O3 and asthma hospital admissions (i.e., both 

10 general and ICU for all ages) is linear. When comparing the curve to a linear fit line, the authors 

11 found that the linear fit was a reasonable approximation of the concentration-response 

12 relationship between O3 and asthma hospital admissions, but the limited data density at relatively 

13 low O3 concentrations contributes to uncertainty in the shape of the concentration-response 

14 relationship at the low end of the distribution of O3 concentrations (U.S. EPA, 2013, Figure 6-

16). 15 

16 In contrast to COPD and asthma, the evidence for pneumonia-related admissions was less 

17 consistent. Medina-Ramon et al. (2006) examined the association between short-term ambient O3

18 concentrations and Medicare hospital admissions among individuals ≥ 65 years of age for 

19 pneumonia. The authors reported an increase in pneumonia hospital admissions in the warm 

20 season for a 30 ppb increase in 8-hour max O3 concentrations, with no evidence of an association 

21 in the cool season or year round. In two-pollutant models restricted to days for which PM10 data 

22 was available, the association between O3 exposure and pneumonia hospital admissions 

23 remained robust. In contrast, Zanobetti and Schwartz (2006) reported a decrease in pneumonia 

24 admissions for a 20 ppb increase in 24-hour average O3 concentrations in Boston for the average 

25 of lags 0 and 1 day. 

26 The magnitude of associations with respiratory-related hospital admissions may be 

27 underestimated due to behavioral modification in response to forecasted air quality (U.S. EPA, 

28 2013, section 4.6.6). Recent studies (Neidell and Kinney, 2010; Neidell, 2009) conducted in 

29 Southern California demonstrates that controlling for avoidance behavior increases O3 effect 

30 estimates for respiratory hospital admissions, specifically for children and older adults. This 

31 study shows that on days where no public alert warning of high O3 concentrations was issued, 

32 there was an increase in asthma hospital admissions. Although only one study has examined 

33 averting behavior and this study is limited to the outcome of asthma hospital admissions in one 

34 location and time period (i.e., Los Angeles, CA for the years 1989-1997), it does provide 

35 preliminary evidence indicating that some epidemiologic studies may underestimate associations 
3B-4 

1 between O3 and health effects by not accounting for behavioral modification when public health 

2 alerts are issued. 

3 Emergency Department Visits for All Respiratory Causes

4 A large single-city study conducted in Atlanta by Tolbert et al. (2007), and subsequently 

5 reanalyzed by Darrow et al. (2011) using different air quality data and evaluating associations 

6 with different metrics, provides evidence for associations between short-term exposures to 

7 ambient O3 concentrations and respiratory emergency department visits. Tolbert et al. (2007) 

8 reported an increase in respiratory emergency department visits for a 30 ppb increase in 8-hour 

9 max O3 concentrations during the warm season. In copollutant models with CO, NO2, and PM10, 

10 limited to days in which data for all pollutants were available, associations between O3 and 

11 respiratory emergency department visits remained positive, but were attenuated. Darrow et al. 

12 (2011) reported the strongest associations with respiratory emergency department visits for 8-

13 hour daily max, 1-hour daily max, and day-time O3 exposure metrics (all associations positive 

14 and statistically significant), while positive, but statistically non-significant, associations were 

15 reported with 24-hour average and commuting period exposure metrics. In addition, a negative 

16 association was observed when using the night-time exposure metric (U.S. EPA, 2013, Figure 6-

17 17). The results of Darrow et al. (2011) suggest that averaging over nighttime hours may lead to 

18 smaller O3 effect estimates for respiratory emergency department visits due to dilution of 

19 relevant O3 concentrations (i.e., the higher concentrations that occur during the daytime); and 

20 potential negative confounding by other pollutants (e.g., CO, NO2) during the nighttime hours 

21 (U.S. EPA, 2013, section 6.2.7.3) 

22 Cause-Specific Emergency Department Visits

23 In evaluating asthma emergency department visits in an all-year analysis, a Canadian 

24 multi-city study (Stieb et al., 2009) reported that 24-hour O3 concentrations were positively 

25 associated with emergency department visits for asthma at lag 1 and lag 2. Though the authors 

26 did not present seasonal analyses, they stated that no associations were observed with emergency 

27 department visits in the winter season, suggesting that the positive associations reported in the 

28 all-year analysis were due to the warm season (Stieb et al., 2009). In addition to asthma, the 

29 authors reported that O3 was positively associated with COPD emergency department visits in 

30 all-year analyses, but that associations with COPD visits were statistically significant only for the 

31 warm season (i.e., April-September). 
3B-5 

1 Several single-city studies have also provided evidence for positive associations between 

2 asthma emergency department visits and ambient O3 concentrations. Ito et al. (2007) reported 

3 positive and statistically significant associations with asthma emergency department visits in 

4 New York City during the warm season, and an inverse association in the cool season, for a 30 

5 ppb increase in 8-hour max O3 concentrations. In two-pollutant models with PM2.5, NO2, SO2, 

6 and CO, the authors found that O3 risk estimates were not substantially changed during the warm 

7 season (U.S. EPA, 2013, Figure 6-20; Table 6-29). 

8 Strickland et al. (2010) examined the association between O3 exposure and pediatric 

9 asthma emergency department visits (ages 5-17 years) in Atlanta using air quality data over the 

10 same years as Darrow et al. (2011) and Tolbert et al. (2007), but using population-weighting to 

11 combine daily pollutant concentrations across monitors. Strickland et al. (2010) reported an 

12 increase in emergency department visits for a 30 ppb increase in 8-hour max O3 concentrations 

13 in an all-year analysis. In seasonal analyses, stronger associations were observed during the 

14 warm season (i.e., May-October) than the cold season. In co-pollutant analyses that included CO, 

15 NO2, PM2.5 elemental carbon, or PM2.5 sulfate, Strickland et al. (2010) reported that O3 risk 

16 estimates were not substantially changed. The authors also examined the concentration-response 

17 relationship between O3 exposure and pediatric asthma emergency department visits and 

18 reported that positive associations with O3 persist at 8-hour ambient O3 concentrations (3-day

19 average of 8-hour daily max concentrations) at least as low as 30 ppb. 

20 In a single-city study conducted in Seattle, WA, Mar and Koenig (2009) examined the 

21 association between O3 exposure and asthma emergency department visits for children (< 18) 

22 and adults (≥ 18). For children, positive and statistically significant associations were reported 

23 across multiple lags, with the strongest associations observed at lag 0 and lag 3. Ozone was also 

24 found to be positively associated with asthma emergency department visits for adults at all lags, 

25 except at lag 0. The slightly different lag times for children and adults suggest that children may 

26 be more immediately responsive to O3 exposures than adults (Mar and Koenig, 2009). 

27 In addition to the U.S. single-city studies discussed above, a single-city study conducted 

28 in Alberta, Canada (Villeneuve et al., 2007) provides support for the findings from Stieb et al. 

29 (2009), but also attempts to identify those lifestages at greatest risk for O3-associated asthma 

30 emergency department visits. Villeneuve et al. reported an increase in asthma emergency 

31 department visits in an all-year analysis across all ages with associations being stronger during 
3B-6 

1 the warmer months. When stratified by age, the strongest associations were observed in the 

2 warm season for individuals 5-14 and 15-44. These associations were not found to be 

3 confounded by the inclusion of aeroallergens in age-specific models.

4 3.2 REFERENCES

5 Biggeri, A; Baccini, M; Bellini, P; Terracini, B. (2005). Meta-analysis of the Italian studies of short-term effects of 

6 air pollution (MISA), 1990-1999. Int J Occup Environ Health 11: 107-122.

7 Cakmak, S; Dales, RE; Judek, S. (2006b). Respiratory health effects of air pollution gases: Modification by 

8 education and income. Arch Environ Occup Health 61: 5-10. http://dx.doi.org/10.3200/AEOH.61.1.5-10

9 Dales, RE; Cakmak, S; Doiron, MS. (2006). Gaseous air pollutants and hospitalization for respiratory disease in the 

10 neonatal period. Environ Health Perspect 114: 1751-1754. http://dx.doi.org/10.1289/ehp.9044

11 Darrow, LA; Klein, M; Sarnat, JA; Mulholland, JA; Strickland, MJ; Sarnat, SE; Russell, AG; Tolbert, PE. (2011). 

12 The use of alternative pollutant metrics in time-series studies of ambient air pollution and respiratory 

13 emergency department visits. J Expo Sci Environ Epidemiol 21: 10-19.

14 .Ito, K; Thurston, GD; Silverman, RA. (2007b). Characterization of PM2.5, gaseous pollutants, and meteorological 

15 interactions in the context of time-series health effects models. J Expo Sci Environ Epidemiol 17: S45-S60.

16 Katsouyanni, K; Samet, JM; Anderson, HR; Atkinson, R; Le Tertre, A; Medina, S; Samoli, E; Touloumi, G; 

17 Burnett, RT; Krewski, D; Ramsay, T; Dominici, F; Peng, RD; Schwartz, J; Zanobetti, A. (2009). Air 

18 pollution and health: A European and North American approach (APHENA). (Research Report 142). 

19 Boston, MA: Health Effects Institute. http://pubs.healtheffects.org/view.php?id=327 

20 Lin, S; Bell, EM; Liu, W; Walker, RJ; Kim, NK; Hwang, SA. (2008a). Ambient ozone concentration and hospital 

21 admissions due to childhood respiratory diseases in New York State, 1991-2001. Environ Res 108: 42-47. 

22 http://dx.doi.org/10.1016/j.envres.2008.06.007

23 Mar, TF; Koenig, JQ. (2009). Relationship between visits to emergency departments for asthma and ozone exposure 

24 in greater Seattle, Washington. Ann Allergy Asthma Immunol 103: 474-479.

25 Medina-Ramon, M; Zanobetti, A; Schwartz, J. (2006). The effect of ozone and PM10 on hospital admissions for 

26 pneumonia and chronic obstructive pulmonary disease: A national multicity study. Am J Epidemiol 163: 

27 579-588. http://dx.doi.org/10.1093/aje/kwj078 

28 Neidell, M. (2009). Information, avoidance behavior, and health: The effect of ozone on asthma hospitalizations. 

29 Journal of Human Resources 44: 450-478. 

30 Neidell, M; Kinney, PL. (2010). Estimates of the association between ozone and asthma hospitalizations that 

31 account for behavioral responses to air quality information. Environ Sci Pol 13: 97-103. 

32 http://dx.doi.org/10.1016/j.envsci.2009.12.006

33 Silverman, RA; Ito, K. (2010). Age-related association of fine particles and ozone with severe acute asthma in New 

34 York City. J Allergy Clin Immunol 125: 367-373. http://dx.doi.org/10.1016/j.jaci.2009.10.061 

35 Stieb, DM; Szyszkowicz, M; Rowe, BH; Leech, JA. (2009). Air pollution and emergency department visits for 

36 cardiac and respiratory conditions: A multi-city time-series analysis. Environ Health Global Access Sci 

37 Source 8: 25. http://dx.doi.org/10.1186/1476-069X-8-25 
3B-7 

1 Strickland, MJ; Darrow, LA; Klein, M; Flanders, WD; Sarnat, JA; Waller, LA; Sarnat, SE; Mulholland, JA; Tolbert, 

2 PE. (2010). Short-term associations between ambient air pollutants and pediatric asthma emergency 

3 department visits. Am J Respir Crit Care Med 182: 307-316. http://dx.doi.org/10.1164/rccm.200908-

4 1201OC 

5 Tolbert, PE; Klein, M; Peel, JL; Sarnat, SE; Sarnat, JA. (2007). Multipollutant modeling issues in a study of ambient 

6 air quality and emergency department visits in Atlanta. J Expo Sci Environ Epidemiol 17: S29-S35. 

7 http://dx.doi.org/10.1038/sj.jes.7500625 

8 U.S. EPA (2013). Integrated Science Assessment of Ozone and Related Photochemical Oxidants (Final Report). 

9 U.S. Environmental Protection Agency, Washington, DC. EPA/600/R-10/076F. Available at: 

10 http://www.epa.gov/ttn/naaqs/standards/ozone/s_o3_2008_isa.html 

11 Villeneuve, PJ; Chen, L; Rowe, BH; Coates, F. (2007). Outdoor air pollution and emergency department visits for 

12 asthma among children and adults: A case-crossover study in northern Alberta, Canada. Environ Health 

13 Global Access Sci Source 6: 40. http://dx.doi.org/10.1186/1476-069X-6-40

14 Wong, CM; Yang, L; Thach, TQ; Chau, PY; Chan, KP; Thomas, GN; Lam, TH; Wong, TW; Hedley, AJ; Peiris, JS. 

15 (2009). Modification by influenza on health effects of air pollution in Hong Kong. Environ Health Perspect 

16 117: 248-253. http://dx.doi.org/10.1289/ehp.11605

17 Yang, Q; Chen, Y; Krewski, D; Burnett, RT; Shi, Y; Mcgrail, KM. (2005). Effect of short-term exposure to low 

18 levels of gaseous pollutants on chronic obstructive pulmonary disease hospitalizations. Environ Res 99: 99-

19 105. http://dx.doi.org/10.1016/j.envres.2004.09.014 

20 Zanobetti, A; Schwartz, J. (2006). Air pollution and emergency admissions in Boston, MA. J Epidemiol Community 

21 Health 60: 890-895. http://dx.doi.org/10.1136/jech.2005.039834
3C-1 

1 APPENDIX 3C

2 AT-RISK POPULATIONS

3 People with Specific Genetic Variants

4 Overall, for variants in multiple genes there is adequate evidence for involvement in 

5 populations being more at-risk than others to the effects of O3 exposure on health (U.S. EPA, 

6 2013, section 8.1). Controlled human exposure and epidemiologic studies have reported evidence 

7 of O3-related increases in respiratory symptoms or decreases in lung function with variants 

8 including GSTM1, GSTP1, HMOX1, and NQO1. NQO1 deficient mice were found to be 

9 resistant to O3-induced AHR and inflammation, providing biological plausibility for results of 

10 studies in humans. Additionally, studies of rodents have identified a number of other genes that 

11 may affect O3-related health outcomes, including genes related to innate immune signaling and 

12 pro- and anti-inflammatory genes, which have not been investigated in human studies. 

13 People with Asthma

14 Previous O3 AQCDs identified individuals with asthma as a population at increased risk 

15 of O3-related health effects. Multiple new epidemiologic studies included in the ISA have 

16 evaluated the potential for increased risk of O3-related health effects in people with asthma, 

17 including: lung function; symptoms; medication use; airway hyperresponsiveness (AHR); and 

18 airway inflammation (also measured as exhaled nitric oxide fraction, or FeNO). A study of 

19 lifeguards in Texas reported decreased lung function with short-term O3 exposure among both 

20 individuals with and without asthma, however, the decrease was greater among those with 

21 asthma (Thaller et al., 2008). A Mexican study of children ages 6-14 detected an association 

22 between short-term O3 exposure and wheeze, cough, and bronchodilator use among asthmatics 

23 but not non-asthmatics, although this may have been the result of a small non-asthmatic 

24 population (Escamilla-Nuñez et al., 2008). A study of modification by AHR (an obligate 

25 condition among asthmatics) reported greater short-term O3-associated decreases in lung 

26 function in elderly individuals with AHR, especially among those who were obese (Alexeeff et 

27 al., 2007). With respect to airway inflammation, in one study, a positive association was reported 

28 for airway inflammation among asthmatic children following short-term O3 exposure, but the 

29 observed association was similar in magnitude to that of non-asthmatics (Barraza-Villarreal et 

30 al., 2008). Similarly, another study of children in California reported an association between O3 

31 concentration and FeNO that persisted both among children with and without asthma as well as 

32 those with and without respiratory allergy (Berhane et al., 2011). Finally, Khatri et al. (2009) 

33 found no association between short-term O3 exposure and altered lung function for either 
3C-2 

1 asthmatic or non-asthmatic adults, but did note a decrease in lung function among individuals 

2 with allergies. 

3 New evidence for difference in effects among asthmatics has been observed in studies 

4 that examined the association between O3 exposure and altered lung function by asthma 

5 medication use. A study of children with asthma living in Detroit reported a greater association 

6 between short-term O3 and lung function for corticosteroid users compared with 

7 noncorticosteroid users (Lewis et al., 2005). Conversely, another study found decreased lung 

8 function among noncorticosteroid users compared to users, although in this study, a large 

9 proportion of non-users were considered to be persistent asthmatics (Hernández-Cadena et al., 

10 2009). Lung function was not related to short-term O3 exposure among corticosteroid users and 

11 non-users in a study taking place during the winter months in Canada (Liu et al., 2009). 

12 Additionally, a study of airway inflammation reported a counterintuitive inverse association with 

13 O3 of similar magnitude for all groups of corticosteroid users and non-users (Qian et al., 2009). 

14 Controlled human exposure studies that have examined the effects of O3 on adults with 

15 asthma and healthy controls are limited. Based on studies reviewed in the 1996 and 2006 O3

16 AQCDs, subjects with asthma appeared to be more sensitive to acute effects of O3 in terms of 

17 FEV1 and inflammatory responses than healthy non-asthmatic subjects. For instance, Horstman 

18 et al. (1995) observed that mild-to-moderate asthmatics, on average, experienced double the 

19 O3-induced FEV1 decrement of healthy subjects (19% versus 10%, respectively, p = 0.04). 

20 Moreover, a statistically significant positive correlation between FEV1 responses to O3 exposure 

21 and baseline lung function was observed in individuals with asthma, i.e., responses increased 

22 with severity of disease. Minimal evidence exists suggesting that individuals with asthma have 

23 smaller O3-induced FEV1 decrements than healthy subjects (3% versus 8%, respectively) 

24 (Mudway et al., 2001). However, the asthmatics in that study also tended to be older than the 

25 healthy subjects, which could partially explain their lesser response since FEV1 responses to O3

26 exposure diminish with age. Individuals with asthma also had significantly more neutrophils in 

27 the BALF (18 hours postexposure) than similarly exposed healthy individuals (Peden et al., 

28 1997; Scannell et al., 1996; Basha et al., 1994). Furthermore, a study examining the effects of O3

29 on individuals with atopic asthma and healthy controls reported that greater numbers of 

30 neutrophils, higher levels of cytokines and hyaluronan, and greater expression of macrophage 

31 cell-surface markers were observed in induced sputum of atopic asthmatics compared with 

32 healthy controls (Hernandez et al., 2010). Differences in O3-induced epithelial cytokine 

33 expression were noted in bronchial biopsy samples from asthmatics and healthy controls (Bosson 

34 et al., 2003). Cell-surface marker and cytokine expression results, and the presence of 

35 hyaluronan, are consistent with O3 having greater effects on innate and adaptive immunity in 

36 these asthmatic individuals. In addition, studies have demonstrated that O3 exposure leads to 
3C-3 

1 increased bronchial reactivity to inhaled allergens in mild allergic asthmatics (Kehrl et al., 1999; 

2 Jorres et al., 1996) and to the influx of eosinophils in individuals with pre-existing allergic 

3 disease (Vagaggini et al., 2002; Peden et al., 1995). Taken together, these results point to several 

4 mechanistic pathways which could account for the enhanced sensitivity to O3 in subjects with 

5 asthma (see Section 5.4.2.2 in the ISA).

6 Toxicological studies provide additional evidence of the biological basis for the greater 

7 effects of O3 among those with asthma or AHR (U.S. EPA, 2013, section 8.2.2). In animal 

8 toxicological studies, an asthmatic phenotype is modeled by allergic sensitization of the 

9 respiratory tract. Many of the studies that provide evidence that O3 exposure is an inducer of 

10 AHR and remodeling utilize these types of animal models. For example, a series of experiments 

11 in infant rhesus monkeys have shown these effects, but only in monkeys sensitized to house dust 

12 mite allergen. Similarly, adverse changes in pulmonary function were demonstrated in mice 

13 exposed to O3; enhanced inflammatory responses were in rats exposed to O3, but only in animals 

14 sensitized to allergen. In general, it is the combined effects of O3 and allergic sensitization which 

15 result in measurable effects on pulmonary function. In a pulmonary fibrosis model, exposure O3

16 for 5 days increased pulmonary inflammation and fibrosis, along with the frequency of 

17 bronchopneumonia in rats. Thus, short-term exposure to O3 may enhance damage in a previously 

18 injured lung (U.S. EPA, 2013, section 8.2.2). 

19 In the 2006 O3 AQCD, the potential for individuals with asthma to have greater risk of 

20 O3-related health effects was supported by a number of controlled human exposure studies, 

21 evidence from toxicological studies, and a limited number of epidemiologic studies. In section 

22 8.2.2, the ISA reports that in the recent epidemiologic literature some, but not all, studies report 

23 greater risk of health effects among individuals with asthma. Studies examining effect measure 

24 modification of the relationship between short-term O3 exposure and altered lung function by 

25 corticosteroid use provided limited evidence of O3-related health effects. However, recent studies 

26 of behavioral responses have found that studies do not take into account individual behavioral 

27 adaptations to forecasted air pollution levels (such as avoidance and reduced time outdoors), 

28 which may underestimate the observed associations in studies that examined the effect of O3

29 exposure on respiratory health (Neidell and Kinney, 2010). This could explain some 

30 inconsistency observed among recent epidemiologic studies. The evidence from controlled 

31 human exposure studies provides support for increased detriments in FEV1 and greater 

32 inflammatory responses to O3 in individuals with asthma than in healthy individuals without a 

33 history of asthma. The collective evidence for increased risk of O3-related health effects among 

34 individuals with asthma from controlled human exposure studies is supported by recent 

35 toxicological studies which provide biological plausibility for heightened risk of asthmatics to 
3C-4 

1 respiratory effects due to O3 exposure. Overall, the ISA finds there is adequate evidence for 

2 asthmatics to be an at-risk population. 

3 Children

4 Children are considered to be at greater risk from O3 exposure because their respiratory 

5 systems undergo lung growth until about 18-20 years of age and are therefore thought to be 

6 intrinsically more at risk for O3-induced damage (U.S. EPA, 2006b). It is generally recognized 

7 that children spend more time outdoors than adults, and therefore would be expected to have 

8 higher exposure to O3 than adults. The ventilation rates also vary between children and adults, 

9 particularly during moderate/heavy activity. Children aged 11 years and older and adults have 

10 higher absolute ventilation rates than children aged 1-11 years. However, children have higher 

11 ventilation rates relative to their lung volumes, which tends to increase dose normalized to lung 

12 surface area. Exercise intensity has a substantial effect on ventilation rate, with high intensity 

13 activities resulting in nearly double the ventilation rate during moderate activity among children 

14 and those adults less than 31 years of age. For more information on time spent outdoors and 

15 ventilation rate differences by age group, see Section 4.4.1 in the ISA (U.S. EPA, 2013). 

16 The 1996 O3 AQCD reported clinical evidence that children, adolescents, and young 

17 adults (<18 years of age) appear, on average, to have nearly equivalent spirometric responses to 

18 O3 exposure, but have greater responses than middle-aged and older adults (U.S. EPA, 1996a). 

19 Symptomatic responses (e.g., cough, shortness of breath, pain on deep inspiration) to O3

20 exposure, however, appear to increase with age until early adulthood and then gradually decrease 

21 with increasing age (U.S. EPA, 1996). Complete lung growth and development is not achieved 

22 until 18-20 years of age in women and the early 20s for men; pulmonary function is at its 

23 maximum during this time as well. 

24 Recent epidemiologic studies have examined different age groups and their risk to 

25 O3-related respiratory hospital admissions and emergency department (ED) visits. Evidence for 

26 greater risk in children was reported in several studies. A study in Cyprus of short-term O3

27 concentrations and respiratory hospital admissions (HA) detected possible effect measure 

28 modification by age with a larger association among individuals < 15 years of age compared 

29 with those > 15 years of age; the effect was apparent only with a 2-day lag (Middleton et al., 

30 2008). Similarly, a Canadian study of asthma-ED visits reported the strongest O3-related 

31 associations among 5- to 14-year olds compared to the other age groups (ages examined 0-75+) 

32 (Villeneuve et al., 2007). Greater O3-associated risk in asthma-related ED visits were also 

33 reported among children (<15 years) as compared to adults (15 to 64 years) in a study from 

34 Finland (Halonen et al., 2009). A study of New York City hospital admissions demonstrated an 

35 increase in the association between O3 exposure and asthma-related hospital admissions for 6- to 
3C-5 

1 18-year olds compared to those < 6 years old and those > 18 years old (Silverman and Ito, 2010). 

2 When examining long-term O3 exposure and asthma HA among children, associations were 

3 determined to be larger among children 1 to 2 years old compared to children 2 to 6 years old 

4 (Lin et al., 2008). A few studies reported positive associations among both children and adults 

5 and no modification of the effect by age. 

6 The evidence reported in epidemiologic studies is supported by recent toxicological 

7 studies which observed O3-induced health effects in immature animals. Early life exposures of 

8 multiple species of laboratory animals, including infant monkeys, resulted in changes in 

9 conducting airways at the cellular, functional, ultra-structural, and morphological levels. The 

10 studies conducted on infant monkeys are most relevant for assessing effects in children. Carey et 

11 al. (2007) conducted a study of O3 exposure in infant rhesus macaques, whose respiratory tract 

12 closely resemble that of humans. Monkeys were exposed either acutely or in episodes designed 

13 to mimic human exposure. All monkeys acutely exposed to O3 had moderate to marked 

14 necrotizing rhinitis, with focal regions of epithelial exfoliation, numerous infiltrating neutrophils, 

15 and some eosinophils. The distribution, character, and severity of lesions in episodically exposed 

16 infant monkeys were similar to that of acutely exposed animals. Neither exposure protocol for 

17 the infant monkeys produced mucous cell metaplasia proximal to the lesions, an adaptation 

18 observed in adult monkeys exposed in another study (Harkema et al., 1987). Functional and 

19 cellular changes in conducting airways were common manifestations of exposure to O3 among 

20 both the adult and infant monkeys (Plopper et al., 2007). In addition, the lung structure of the 

21 conducting airways in the infant monkeys was significantly stunted by O3 and this aberrant 

22 development was persistent 6 months postexposure (Fanucchi et al., 2006). 

23 Age may also affect the inflammatory response to O3 exposure. Toxicological studies 

24 reported that the difference in effects among younger lifestage test animals may be due to 

25 age-related changes in antioxidants levels and sensitivity to oxidative stress. Further discussion 

26 of these studies may be found in section 8.3.1.1 of the ISA (U.S. EPA, 2013, p. 8-18). 

27 The previous and recent human clinical and toxicological studies reported evidence of 

28 increased risk from O3 exposure for younger ages, which provides coherence and biological 

29 plausibility for the findings from epidemiologic studies. Although there was some inconsistency, 

30 generally, the epidemiologic studies reported positive associations among both children and 

31 adults or just among children. The interpretation of these studies is limited by the lack of 

32 consistency in comparison age groups and outcomes examined. However, overall, the 

33 epidemiologic, controlled human exposure, and toxicological studies provide adequate evidence 

34 that children are potentially at increased risk of O3-related health effects.
3C-6 

1 Older Adults

2 The ISA notes that older adults are at greater risk of health effects associated with O3

3 exposure through a variety of intrinsic pathways (U.S. EPA, 2013, section 8.3.1.2). In addition, 

4 older adults may differ in their exposure and internal dose. Older adults were outdoors for a 

5 slightly longer proportion of the day than adults aged 18-64 years. Older adults also have 

6 somewhat lower ventilation rates than adults aged 31 - less than 61 years. For more information 

7 on time spent outdoors and ventilation rate differences by age group, see Section 4.4 in the ISA 

8 (U.S. EPA, 2013). The gradual decline in physiological processes that occur with aging may lead 

9 to increased risk of O3-related health effects (U.S. EPA, 2006a). Respiratory symptom responses 

10 to O3 exposure appears to increase with age until early adulthood and then gradually decrease 

11 with increasing age (U.S. EPA, 1996a); lung function responses to O3 exposure also decline 

12 from early adulthood (U.S. EPA, 1996a). The reductions of these responses with age may put 

13 older adults at increased risk for continued O3 exposure. In addition, older adults, in general, 

14 have a higher prevalence of preexisting diseases compared to younger age groups and this may 

15 also lead to increased risk of O3-related health effects (U.S. EPA, 2013, section 8.3.1.2). With 

16 the number of older Americans increasing in upcoming years (estimated to increase from 12.4% 

17 of the U.S. population to 19.7% between 2000 to 2030, which is approximately 35 million and 

18 71.5 million individuals, respectively) this group represents a large population potentially at risk 

19 of O3-related health effects (SSDAN CensusScope, 2010a; U.S. Census Bureau, 2010). 

20 The majority of recent studies reported greater effects of short-term O3 exposure and 

21 mortality among older adults, which is consistent with the findings of the 2006 O3 AQCD. A 

22 study (Medina-Ramón and Schwartz, 2008) conducted in 48 cities across the U.S. reported larger 

23 effects among adults ≥65 years old compared to those < 65 years; further investigation of this 

24 study population revealed a trend of O3-related mortality risk that gets larger with increasing age 

25 starting at age (Zanobetti and Schwartz, 2008). Another study conducted in 7 urban centers in 

26 Chile reported similar results, with greater effects in adults ≥65 years old (Cakmak et al., 2007). 

27 More recently, a study conducted in the same area reported similar associations between O3

28 exposure and mortality in adults aged < 64 years old and 65 to 74 years old, but the risk was 

29 increased among older age groups (Cakmak et al., 2011). A study performed in China reported 

30 greater effects in populations ≥45 years old (compared to 5 to 44 year olds), with statistically 

31 significant effects present only among those ≥65 years old (Kan et al., 2008). An Italian study 

32 reported higher risk of all-cause mortality associated with increased O3 concentrations among 

33 individuals ≥85 year old as compared to those 35 to 84 years old (Stafoggia et al., 2010). The Air 

34 Pollution and Health: A European and North American Approach (APHENA) project examined 

35 the association between O3 exposure and mortality for those <75 and ≥ 75 years of age. In 
3C-7 

1 Canada, the associations for all-cause and cardiovascular mortality were greater among those 

2 ≥75 years old. In the U.S., the association for all-cause mortality was slightly greater for those 

3 <75 years of age compared to those ≥75 years old in summer-only analyses. No consistent 

4 pattern was observed for CVD mortality. In Europe, slightly larger associations for all-cause 

5 mortality were observed in those <75 years old in all-year and summer-only analyses. Larger 

6 associations were reported among those <75years for CVD mortality in all-year analyses, but the 

7 reverse was true for summer-only analyses (Katsouyanni et al., 2009). 

8 With respect to epidemiologic studies of O3 exposure and hospital admissions, a positive 

9 association was reported between short-term O3 exposure and respiratory hospital admissions for 

10 adults ≥65 years old but not for those adults aged 15 to 64 years (Halonen et al., 2009). In the 

11 same study, no association was observed between O3 concentration and respiratory mortality 

12 among those ≥65 years old or those 15 to 64 years old. No modification by age (40 to 64 year 

13 olds versus >64 year olds) was observed in a study from Brazil examining O3 levels and COPD 

14 ED visits.

15 Although some outcomes reported mixed findings regarding an increase in risk for older 

16 adults, recent epidemiologic studies report consistent positive associations between short-term 

17 O3 exposure and mortality in older adults. The evidence from mortality studies is consistent with 

18 the results reported in the 2006 O3 AQCD and is supported by toxicological studies providing 

19 biological plausibility for increased risk of effects in older adults. Also, older adults may be 

20 experiencing increased exposure compared to younger adults. Overall, the ISA (U.S. EPA 2013) 

21 concludes adequate evidence is available indicating that older adults are at increased risk of 

22 O3-related health effects. 

23 People with Diets Lower in Vitamins C and E

24 Diet was not examined as a factor potentially affecting risk in previous O3 AQCDs, but 

25 recent studies have examined modification of the association between O3 and health effects by 

26 dietary factors. Because O3 mediates some of its toxic effects through oxidative stress, the 

27 antioxidant status of an individual is an important factor that may contribute to increased risk of 

28 O3-related health effects. Supplementation with vitamins C and E has been investigated in a 

29 number of studies as a means of inhibiting O3-mediated damage. 

30 Two epidemiologic studies have examined effect measure modification by diet and found 

31 evidence that certain dietary components are related to the effect O3 has on respiratory outcomes. 

32 In one recent study the effects of fruit/vegetable intake and Mediterranean diet were examined. 

33 Increases in these food patterns, which have been noted for their high vitamins C and E and 

34 omega-3 fatty acid content, were positively related to lung function in asthmatic children living 
3C-8 

1 in Mexico City, and modified by O3 exposure (Romieu et al., 2009). Another study examined 

2 supplementation of the diets of asthmatic children in Mexico with vitamins C and E (Sienra3 Monge et al., 2004). Associations were detected between short-term O3 exposure and nasal 

4 airway inflammation among children in the placebo group but not in those receiving the 

5 supplementation. 

6 The epidemiologic evidence is supported by controlled human exposure studies, 

7 discussed in section 8.4.1 of the ISA (U.S. EPA 2013), that have shown that the first line of 

8 defense against oxidative stress is antioxidants-rich extracellular lining fluid (ELF) which 

9 scavenge free radicals and limit lipid peroxidation. Exposure to O3 depletes antioxidant levels in 

10 nasal ELF probably due to scrubbing of O3; however, the concentration and the activity of 

11 antioxidant enzymes either in ELF or plasma do not appear to be related to O3 responsiveness. 

12 Controlled studies of dietary antioxidant supplementation have demonstrated some protective 

13 effects of α-tocopherol (a form of vitamin E) and ascorbate (vitamin C) on spirometric measures 

14 of lung function after O3 exposure but not on the intensity of subjective symptoms and 

15 inflammatory responses. Dietary antioxidants have also afforded partial protection to asthmatics 

16 by attenuating postexposure bronchial hyperresponsiveness. Toxicological studies discussed in 

17 section 8.4.1 of the ISA (U.S. EPA 2013) provide evidence of biological plausibility to the 

18 epidemiologic and controlled human exposure studies. 

19 There is adequate evidence that individuals with diets lower in vitamins C and E are at 

20 risk for O3-related health effects. The evidence from epidemiologic studies is supported by 

21 controlled human exposure and toxicological studies. 

22 Outdoor Workers

23 Studies included in the 2006 O3 AQCD reported that individuals who participate in 

24 outdoor activities or work outside to be a population at increased risk based on consistently 

25 reported associations between O3 exposure and respiratory health outcomes in these groups (U.S. 

26 EPA, 2006b). Outdoor workers are exposed to ambient O3 concentrations for a greater period of 

27 time than individuals who spend their days indoors. As discussed in Section 4.7 of the ISA (U.S. 

28 EPA, 2013) outdoor workers sampled during the work shift had a higher ratio of personal 

29 exposure to fixed-site monitor concentrations than health clinic workers who spent most of their 

30 time indoors. Additionally, an increase in dose to the lower airways is possible during outdoor 

31 exercise due to both increases in the amount of air breathed (i.e., minute ventilation) and a shift 

32 from nasal to oronasal breathing. The association between FEV1 responses to O3 exposure and 

33 minute ventilation is discussed more fully in Section 6.2.3.1 of the 2006 O3 AQCD. 
3C-9 

1 Previous studies have shown that increased exposure to O3 due to outdoor work leads to 

2 increased risk of O3-related health effects, specifically decrements in lung function (U.S. EPA, 

3 2006b). The strong evidence from the 2006 O3 AQCD which demonstrated increased exposure, 

4 dose, and ultimately risk of O3-related health effects in this population supports the conclusion 

5 that there is adequate evidence to indicate that increased exposure to O3 through outdoor work 

6 increases the risk of O3-related health effects.

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24 inflammation in Mexico City school children exposed to air pollutants. Respir Res 10: 122.

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30 nasal inflammatory responses among young asthmatics exposed to high levels of ozone. Clin Exp Immunol 

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3C-12 

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3 http://www.epa.gov/ttn/naaqs/standards/ozone/s_o3_2008_isa.html 

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9 assessment [EPA Report]. (EPA/600/P-03/004A). Washington, DC. 

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12 Environmental Protection Agency, Washington, DC. EPA/600/R-05/004aF-cF. March 2006. Available at: 

13 http://www.epa.gov/ttn/naaqs/standards/ozone/s_o3_cr_cd.html

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19 Villeneuve, PJ; Chen, L; Rowe, BH; Coates, F. (2007). Outdoor air pollution and emergency department visits for 

20 asthma among children and adults: A case-crossover study in northern Alberta, Canada. Environ Health 

21 Global Access Sci Source 6: 40. http://dx.doi.org/10.1186/1476-069X-6-40

22 Zanobetti, A; Schwartz, J. (2008a). Is there adaptation in the ozone mortality relationship: A multi-city case23 crossover analysis. Environ Health 7: 22. http://dx.doi.org/10.1186/1476-069X-7-22 
3D-1 

1 APPENDIX 3-D: AMBIENT O3 CONCENTRATIONS IN LOCATIONS OF 

2 HEALTH STUDIES 

Annual 4th 3 highest daily maximum O3 concentrations for all U.S. monitors operating 

4 during the 1975 – 2010 period were retrieved from EPA’s AQS database. These data were used 

5 to calculate O3 design values for the 2008 8-hour O3 NAAQS of 0.075 parts per million (ppm) 

6 according to 40 CFR part 50, Appendix P. Design values were calculated for each O3 monitor 

7 and each 3-year period between 1975-1977 and 2008-2010 whenever sufficient data were 

8 available. 

9 Ozone Design Values in Study Locations 

10 Ozone monitors were matched to 200 health study locations on a case-by-case basis, using 

11 the following guidelines: 

12 1) Areas defined by a Metropolitan Statistical Area (MSA) were matched with O3 monitors 

13 by incorporating all of the monitors located in within the MSA boundaries. 

14 2) Areas not represented by a MSA were matched to monitors by incorporating all of the 

15 monitors in the county central to location of the health study area. 

16 3) In some cases, EPA staff made judgment calls. For example, EPA staff matched the Los 

17 Angeles, CA study area to the Los Angeles-Long Beach-Santa Ana, CA MSA defined by 

18 Los Angeles County, CA and Orange County, CA, while the Long Beach, CA study area 

19 was matched to Los Angeles County, CA and the Santa Ana, CA study area was matched 

20 to Orange County, CA. 

21 In some cases, EPA staff matched two or more study areas to the same county or MSA. 

22 In other cases, a study area was matched to a MSA and another study area was matched to a 

23 county within the same MSA. For each 3-year period, the area design value was determined by 

24 the monitor reporting the highest design value in the county or MSA. This has two implications 

25 for the design values: 

26 1) Design values are sensitive to changes in the monitoring network. The addition or 

27 discontinuation of O3 monitors in an area may cause increases or decreases in the design 

28 value trend. 

29 2) Only valid design values are reported. According to 40 CFR Part 50, Appendix P, design 

30 values greater than the level of the NAAQS (0.075 ppm) are always valid, while design 

31 values less than or equal to 0.075 ppm must have 75% annual data completeness in order 

32 to be valid. This may cause anomalies in the design value trend. For example, a monitor 

33 may report a valid design value based on as few as 12 days of data, or a monitor with less 

34 than 75% annual data completeness may have valid design values in some 3-year periods 

35 and invalid design values in others. 
3D-2 

36 We have identified design values for the U.S. O3 epidemiologic studies identified in Sections 

37 3.1.4.2 and 3.1.4.3 of the second draft Policy Assessment (see Tables 3D-1 and 3D-2, 

38 respectively). For each study, design values were identified for the cities evaluated and for the 

39 years over which the study was conducted. These design values are reported in tables A-1 to A40 22 of the Wells et al, 2012 memo “Analysis of Recent U.S. Ozone Air Quality Data to Support 

41 the O3 NAAQS Review and Quadratic Rollback Simulations to Support the First Draft of the 

42 Risk and Exposure Assessment”. 

43 Table 3D-1 Number of Study Cities from Epidemiologic Studies Using Short-Term O3

Metrics with 3-Year Averages of Annual 4th 44 Highest Daily Maximum 8-hour 

O3 Concentrations > 75, 70, 65, or 60 ppb145

Study

Number 

of Cities

Study 

Period

Number 

(Percent) of 

Cities >75

Number 

(Percent) of 

Cities >70

Number 

(Percent) of 

Cities >65

Number 

(Percent) of 

Cities >60

Bell et al. 20042 95 1987-2000 89 (94%) 91 (96%) 93 (98%) 94 (99%)

Bell et al. 2006, 2007, 2008 98 1987-2000 92 (94%) 94 (96%) 96 (98%) 97 (99%)

Cakmak al. 2006 10 1993-2000 3 (30%) 3 (30%) 4 (40%) 8 (80%)

Dales et al. 2006 11 1986-2000 4 (36%) 6 (55%) 7 (64%) 10 (91%)

Franklin & Schwartz, 2008 18 2000-2005 17 (94%) 18 (100%) 18 (100%) 18 (100%)

Katsouyanni et al. 200938941987-1996 83 (93%) 84 (94%) 86 (97%) 88 (99%)

Katsouyanni et al. 2009514 1987-1996 12 (86%) 12 (86%) 14 (100%) 14 (100%)

Katsouyanni et al. 2009612 1987-1996 4 (33%) 6 (50%) 7 (58%) 11 (92%)

Medina-Ramon et al. 2006 2671986-1999 24 (92%) 24 (92%) 26 (100%) 26 (100%)

Schwartz, 2005 14 1986-1993 13 (93%) 13 (93%) 14 (100%) 14 (100%)

Stieb et al. 2009 7 1992-2003 2 (29%) 2 (29%) 3 (43%) 4 (57%)

Zanobetti & Schwartz, 

2008a,b

48 1989-2000 44 (92%) 45 (94%) 47 (98%) 47 (98%)

 

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. 

2We also evaluated the bayes-adjusted effect estimates for individual cities presented by Bell et al. (2004). None of 

the cities for which individual city effect estimates were statistically significant would have met the current standard 

over the study period. 

3

 U.S. cities; examining mortality 

4

 Study authors included 90 cities in their analyses; air quality data that met completeness criteria described above 

were available for 89 cities 

5

 U.S. cities; examining morbidity 

6

 Canadian cities; examining mortality and morbidity 

7

 Study authors included 27 cities in their analyses; air quality data that met completeness criteria described above 

were available for 26 cities
3D-3 

46 Table 3D-2 Number of Study Cities from Epidemiologic Studies Using Long-Term O3

Metrics with 3-Year Averages of Annual 4th 47 Highest Daily Maximum 8-hour 

48 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 1181994-2003 11 (100%) 11 (100%) 11 (100%) 11 (100%)

Jerrett et al. 2009 9491977-2000 91 (97%) 92 (98%) 93 (99%) 94 (100%)

Lin et al. 2008 2610 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 11

11 1992-2005 12 (100%) 12 (100%) 12 (100%) 12 (100%)

Zanobetti & Schwartz 2011 105 1985-2006 100 (95%) 104 (99%) 104 (99%) 104 (99%)

49

50

 

8

 Study authors included 12 cities in their analyses, air quality data that met completeness criteria described above 

were available for 11 cities 

9

 Study authors included 96 cities in their analyses, air quality data that met completeness criteria described above 

were available for 94 cities 

10 Study authors included 27 cities in their analyses, air quality data that met completeness criteria described 

above were available for 26 cities 

11 Study authors included 12 cities in their analyses, air quality data that met completeness criteria described 

above were available for 11 cities 
3D-4 

51 Relationship between average 24-hour and highest 8-hour O3 concentrations for cities 

52 analyzed by Bell et al. (2006) 

53 Bell et al. (2006) reported associations between mortality and 24-hour average O3

54 concentrations (i.e., averaged across monitors in cities with multiple monitors) in a multi-city 

55 study of 98 U.S. cities. Positive associations persisted in a series of analyses that restricted O3

56 concentrations to those below various cut points (cut points ranged from 5 to 60 ppb in 5 ppb 

57 increments). To facilitate consideration of these cut point analyses for the second draft of the O3

58 Policy Assessment, so as to match the form and averaging time of the existing primary standard, 

59 we evaluated the relationship between 24-hour average O3 concentrations, averaged across 

60 monitors in cities with multiple monitors, and the highest 8-hour daily maximum O3

61 concentrations among the individual monitors in each city. 

62 EPA staff retrieved daily 24-hour average and 8-hour maximum O3 concentrations 

63 reported to EPA by monitors in the 98 study areas defined in Bell et al. (2006) during the 1987-

64 2012 period from EPA’s Air Quality System (AQS) database. Next, EPA staff obtained the 

65 study area boundaries from the published study (Bell et al., 2006) and used them to determine 

66 which O3 monitoring sites were associated with each study area. The 24-hour average O3

67 concentrations were averaged spatially across all available monitors within each study area on 

68 each day where monitoring data were collected. Next, days where the area-wide 24-hour average 

69 concentration (i.e., averaged spatially across monitors in areas with multiple monitors) was 

70 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 71 highest 8-hour daily maximum 

72 concentrations were identified for each study area and for each year from 1987-2012 (Table 3D73 3). This process was repeated by further removing days with area-wide 24-hour average 

74 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 75 highest 8-hour 

76 daily maximum concentrations for all study areas. 

77
3D-5 

Table 3D-3 Number of Study Cities with 4th 78 Highest 8-hour Daily Maximum 

79 Concentrations Greater Than the Level of the Current Standard and 

80 Potential Alternative Standards For Various Cut-Point Analyses Presented 

in Bell et al. (2006)12 81

82

83

 

12 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)
3D-6 

84 Relationship between average and highest 8-hour daily maximum O3 concentrations for 

85 New York City, as analyzed by Silverman and Ito (2010) 

86 EPA staff retrieved daily maximum 8-hour O3 concentrations for the 13 monitors in the 

87 New York City area used in the Silverman and Ito (2010) study for April-August of 1999-2006 

88 from the AQS database. Next, EPA staff spatially averaged these concentrations across monitors 

89 for each day during this period, and then paired them with the highest 8-hour daily maximum 

90 value reported across the 13 monitors on each day. 

91 Next, the range of observed average daily maximum 8-hour concentrations was broken 

92 into 5 ppb increments. The number of days where the area-wide average daily maximum 8-hour 

93 concentration fell within the increment and the number of days where one or more monitored 8-

94 hour daily maximum values were greater than 75, 70, 65 and 60 ppb were recorded for each 5 

95 ppb increment. These numbers are summarized in Table 3D-4. 

96 Table 3D-4 Summary statistics for Observed O3 Concentrations in the New York City 

97 Area, April – August 1999 – 2006

98

99

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)
3D-7 

100 Relationship between average and highest 8-hour daily maximum O3 concentrations for 

101 Atlanta, as analyzed by Strickland et al. (2010) 

102 For our assessment of the Strickland et al. (2010) study, based in the Atlanta metropolitan 

103 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 104

105 monitor was a part of the Southeastern Aerosol Research and Characterization (SEARCH) 

106 network, which does not report data to EPA. EPA staff calculated the area-wide average of the 

107 8-hour daily maximum concentrations for each day, and compared to population-weighted 

108 average concentrations obtained from the author. The correlation between the arithmetic average 

109 values and the population-weighted average values was very high (R = 0.985), thus EPA staff 

110 deemed the arithmetic average to be a suitable surrogate for the population-weighted average 

111 used in the study. Finally, 3-day moving averages were calculated from the daily area-wide 

112 average values (matching the air quality metric used in the study), and paired with the highest 

113 monitored 8-hour daily maximum value occurring during each 3-day period. 

114 Next, the range of observed average daily maximum 8-hour concentrations was broken 

115 into 5 ppb increments. The number of days where the area-wide average daily maximum 8-hour 

116 concentration fell within the increment and the number of days where one or more monitored 8-

117 hour daily maximum values were greater than 75, 70, 65 and 60 ppb were recorded for each 5 

118 ppb increment. These numbers are summarized in Table 3D-5. 

119 Table 3D-5 Summary statistics for Observed O3 Concentrations in the Atlanta Area, 

120 April – August 1999 – 2006

121

122

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)
3D-8 

123 Relationship between annual and highest 1-hour daily maximum O3 concentrations for 12 

124 study areas, as analyzed by Jerrett et al. (2009) 

125 The Jerrett et al. (2009) study used a long-term metric based on seasonal averages of 1-

126 hour daily maximum O3 concentrations to evaluate associations between respiratory mortality 

127 and long-term or repeated exposures to O3. Authors divided study cities into quartiles based on 

128 these seasonal averages of 1-hour daily O3 concentrations. Using AQS, we identified the 3-year 

averages of annual 4th 129 highest daily maximum 8-hour O3 concentrations in study cities during the 

130 study period. Table 3D-6 presents the means and maximums of these concentrations over the 

131 study period. 

132 In addition, for the 12 urban case study areas included in the epidemiology-based risk 

assessment of the 2nd 133 draft of the Health REA we identified the seasonal averages of 1-hour daily 

134 maximum concentrations (i.e., the O3 metric evaluated by Jerrett et al., 2009) for air quality 

135 adjusted to the current and alternative standards. Specifically, for adjusted air quality “quarterly” 

136 averages of 1-hour concentrations for April-June and July-August were calculated for each area 

137 and year. The quarterly values were considered to be valid if valid daily maximum 1-hour 

138 values were available for at least 75% of the days in the quarter. The two quarterly values were 

139 then averaged, as was done by Jerrett et al. (2009) to generate the long-term metric used in the 

140 study. This process was repeated for the various model-based adjustment scenarios in each of 

141 the 12 study areas. Summary statistics based on this seasonal average of daily O3 concentrations 

142 are presented in Table 3D-7 for recent air quality and for air quality adjusted to just meet the 

143 current and alternative standards. 

144
3D-9 

Table 3D-6 Three-Year Averages of Annual 4th 145 Highest Daily Maximum 8-hour O3

Concentrations in 9413 146 Study Areas Examined in Jerrett et al. (2009) 

City Mean over

study 

period

Max over 

study 

period

Cities in the lowest quartile of average exposure14

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 

exposure15

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

 

13 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). 

14 Based on visual inspection of Figure 1 in Jerrett et al. (2009) 

15 Based on visual inspection of Figure 1 in Jerrett et al. (2009) 
3D-10

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 NA16 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

 

16 Air quality data did not meet completeness criteria described above 
3D-11

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

147 
3D-12

148 Table 3D-7 Long-Term O3 Concentrations in 12 Urban Case Study Areas (Using the O3

149 Metric Evaluated by Jerrett et al., 2009) for Recent Air Quality and Air 

150 Quality Adjusted to Meet Standard Levels of 75, 70, 65, and 60 ppb 

151

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
3D-13

152 REFERENCES 

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154 exposure and mortality in 98 US communities. Am J Epidemiol 167: 986-997. 

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162 on cardiac disease? J Occup Environ Med 48: 89-94. 

163 http://dx.doi.org/10.1097/01.jom.0000184878.11956.4b 

164 Cakmak, S; Dales, RE; Judek, S. (2006b). Respiratory health effects of air pollution gases: Modification by 

165 education and income. Arch Environ Occup Health 61: 5-10. http://dx.doi.org/10.3200/AEOH.61.1.5-10

166 Dales, RE; Cakmak, S; Doiron, MS. (2006). Gaseous air pollutants and hospitalization for respiratory disease in the 

167 neonatal period. Environ Health Perspect 114: 1751-1754. http://dx.doi.org/10.1289/ehp.9044 

168 Franklin, M; Schwartz, J. (2008). The impact of secondary particles on the association between ambient ozone and 

169 mortality. Environ Health Perspect 116: 453-458. http://dx.doi.org/10.1289/ehp.10777 

170 Islam, T; Berhane, K; McConnell, R; Gauderman, WJ; Avol, E; Peters, JM; Gilliland, FD. (2009). Glutathione-S171 transferase (GST) P1, GSTM1, exercise, ozone and asthma incidence in school children. Thorax 64: 197-

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174 and risk of asthma during adolescence. Am J Respir Crit Care Med 177: 388-395. 

175 http://dx.doi.org/10.1164/rccm.200706-863OC 

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186 pneumonia and chronic obstructive pulmonary disease: A national multicity study. Am J Epidemiol 163: 

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190 http://dx.doi.org/10.1136/jech.2008.083576 
3D-14

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210
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

A

Species 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 

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6A-1 

1 APPENDIX 6A

2 

3 Calculation of Approximate Equivalent 12-hr SUM06 and 12-hr W126

4 

5 

6 SOURCE: 2007 Staff Paper, Appendix 7B (U.S. EPA, 2007). 

7 Despite various metrics reported in the vegetation effects literature, there is no standard

8 method for calculating equivalent levels between metrics. The maximum 3-month 12-hr 

9 SUM06 of 25 ppm-hr secondary standard that was proposed in the last review (62 FR 

10 38877) was based on a yield loss prevention of approximately 10% in 50% of crop cases 

11 studied in the National Crop Loss Analysis Network (NCLAN) experiments. For 

12 consistency, staff judged it appropriate to use the NCLAN experiments to derive 

13 equivalents between the 12-hr SUM06 and W126. For example, below are the 12-hr 

14 SUM06 and W126 NCLAN equations to protect 50% of crop cases from a specified

15 percent yield loss (Lee and Hogsett 1996): 

 16 

Metric Weibull Equation

12-hr SUM06 Predicted Relative Yield Loss = 1- exp(-[SUM06/87.42]^1.82)

12-hr W126 Predicted Relative Yield Loss = 1- exp(-[W126/96.05]^1.48)

 17 

18 In the first equation, solving for a SUM06 of 25 ppm-hr equals a predicted relative yield

19 loss of 10%. Solving the second equation for a 10% yield loss equals a W126 of 21 ppmhr. 

20 Thus, staff considers a 12-hr SUM06 of 25 ppm-hr and a 12-hr W126 of 21 ppm-hr 

21 approximately equivalent. 

 22 

23 REFERENCES

24 

25 Lee, E. H.; Hogsett, W. E. (1996) Methodology for calculating inputs for ozone secondary standard benefits 

26 analysis: part II. Report prepared for Office of Air Quality Planning and Standards, Air Quality Strategies 

27 and Standards Division, U.S. Environmental Protection Agency, Research Triangle Park, N.C., March.

28 U.S. EPA (U.S. Environmental Protection Agency). (2007). Review of the national ambient air quality standards for 

29 ozone: Policy assessment of scientific and technical information: OAQPS staff paper [EPA Report]. 

30 (EPA/452/R-07/003). Research Triangle Park, NC. 

31 http://www.epa.gov/ttn/naaqs/standards/ozone/data/2007_01_ozone_staff_paper

32 

33 
United States

Environmental Protection

Agency

Office of Air Quality Planning and Standards

Health and Environmental Impacts Division

Research Triangle Park, NC

Publication No. EPA–452/P–14–002

January 2014