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7 INTEGRATED HEALTH EFFECTS OF LONG-TERM OZONE 

EXPOSURE

7.1 Introduction

1 This chapter reviews, summarizes, and integrates the evidence on relationships between 

2 health effects and long-term exposures to O3. Both epidemiologic and toxicological 

3 studies provide a basis for examining long-term O3 exposure health effects for respiratory 

4 effects, cardiovascular effects, reproductive and developmental effects, central nervous 

5 system effects, cancer outcomes, and mortality. Long-term exposure has been defined as 

a duration of approximately 30 days (1 month) or longer1 6 . However, in order to 

7 characterize the weight of evidence for the effects of O3 on reproductive and 

8 developmental effects in a consistent, cohesive and integrated manner, results from both 

9 short-term and long-term exposure periods are included in that section, and are identified 

10 accordingly in the text and tables.

11 Conclusions from the 2006 O3 AQCD (U.S. EPA, 2006b) are summarized briefly at the 

12 beginning of each section, and the evaluation of evidence from recent studies builds upon 

13 what was available during the previous review. For each health outcome (e.g., respiratory 

14 disease, lung function), results are summarized for studies from the specific scientific 

15 discipline, i.e., epidemiologic and toxicological studies. The major sections 

16 (i.e., respiratory, cardiovascular, mortality, reproductive/developmental, cancer) conclude 

17 with summaries of the evidence for the various health outcomes within that category and 

18 integration of the findings that lead to conclusions regarding causality based upon the 

19 framework described in the Preamble to this ISA. Determination of causality is made for 

20 the overall health effect category, such as respiratory effects, with coherence and 

21 plausibility being based on evidence from across disciplines and also across the suite of 

22 related health outcomes, including cause-specific mortality. 

23 As mentioned in Chapter 2 (Section 2.3), epidemiologic studies generally present O3-

24 related effect estimates for mortality and morbidity health outcomes based on an 

25 incremental change in exposure. Studies traditionally present the relative risk per an 

26 incremental change equal to the interquartile range in O3 concentrations or some other 

27 arbitrary value (e.g., 10 ppb). Additionally, various exposure metrics are used in O3 

28 epidemiologic studies, with the three most common being the maximum 1-h average 

29 within a 24-h period (1-h max), the maximum 8-h average within a 24-h period 

30 (8-h max), and 24-h average (24-h avg). For the purpose of presenting results from 

 1 Unless otherwise specified, the term “chronic” generally refers to an annual exposure duration for epidemiology 

studies and a duration of greater than 10% of the lifespan of the animal in toxicological studies.
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1 studies that use different exposure metrics, EPA consistently applies the same O3

2 increments to facilitate comparisons between the results of various studies that may 

3 present results for different incremental changes. Differences due to the use of varying 

4 exposure metrics (e.g., 1-h max, 24-h avg) become less apparent when averaged across 

5 longer exposure periods, because levels are typically lower and less variable. As such, 

6 throughout this chapter an increment of 10 ppb was consistently applied across studies, 

7 regardless of exposure metric, to facilitate comparisons between the results from these 

8 studies.

7.2 Respiratory Effects

9 Studies reviewed in the 2006 O3 AQCD examined evidence for relationships between 

10 long-term O3 exposure (several months to yearly) and effects on respiratory health 

11 outcomes including declines in lung function, increases in inflammation, and 

12 development of asthma in children and adults. Animal toxicology data provided a clearer 

13 picture indicating that long-term O3 exposure may have lasting effects. Chronic exposure 

14 studies in animals have reported biochemical and morphological changes suggestive of 

15 irreversible long-term O3 impacts on the lung. In contrast to supportive evidence from 

16 chronic animal studies, the epidemiologic studies on longer-term (annual) lung function 

17 declines, inflammation, and new asthma development remained inconclusive. 

18 Several studies reviewed in the 2006 O3 AQCD (Horak et al., 2002; Frischer et al., 1999)

19 collectively indicated that O3 exposure averaged over several summer months was 

20 associated with smaller increases in lung function growth in children. For longer 

21 averaging periods (annual), the definitive analysis in the Children’s Health Study (CHS) 

22 reported by Gauderman et al. (2004) provided little evidence that such long-term 

23 exposure to ambient O3 was associated with significant deficits in the growth rate of lung 

24 function in children in contrast to the effects observed with other pollutants such as acid 

25 vapor, NO2, and PM2.5. Limited epidemiologic research examined the relationship 

26 between long-term O3 exposures and inflammation. Consistent with evidence of 

27 inflammation and allergic responses reported in experimental studies, an association 

28 between 30-day average O3 and increased eosinophil levels was observed in an Austrian 

29 study (Frischer et al., 2001). The cross-sectional studies available for the 2006 O3 AQCD 

30 detected no associations between long-term O3 exposures and asthma prevalence, asthma31 related symptoms or allergy to common aeroallergens in children after controlling for 

32 covariates. However, longitudinal studies provided evidence that long-term O3 exposure 

33 influences the risk of asthma development in children (McConnell et al., 2002) and adults

34 (McDonnell et al., 1999a; Greer et al., 1993). 
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1 New evidence presented below reports interactions between genetic variants and long2 term O3 exposure in effects on new onset asthma in U.S. cohorts in multi-community 

3 studies where protection by specific oxidant gene variants was restricted to children 

4 living in low O3 communities. Related studies report coherent relationships between 

5 respiratory symptoms among asthmatics and long-term O3 exposure. Short-term exposure 

6 to O3 is associated with increases in respiratory symptoms and asthma medication use in 

7 children with asthma (Section 6.2.4.1) and asthma hospitalizations in children 

8 (Section 6.2.7.2). A new line of evidence reports a positive concentration-response 

9 relationship between first asthma hospitalization and long-term O3 exposure. Results 

10 from recent studies examining pulmonary function, inflammation, and allergic responses 

11 are also presented. 

7.2.1 Asthma

7.2.1.1 New Onset Asthma

12 Asthma is a heterogeneous disease with a high degree of temporal variability. Its 

13 progression and symptoms can vary within an individual’s experience over time. The 

14 course of asthma may vary markedly between young children, older children and 

15 adolescents, and adults. This variation is probably more dependent on age than on 

16 symptoms (NHLBI, 2007). Longitudinal cohort studies have examined associations 

17 between long-term O3 exposures and the onset of asthma in adults and children 

18 (McConnell et al., 2002; McDonnell et al., 1999a; Greer et al., 1993), with results 

19 indicating a direct effect of long-term O3 exposure on asthma risk in adults and effect 

20 modification by O3 in children. 

21 Associations between long-term O3 exposure and new cases of asthma were reported in a 

22 cohort of nonsmoking adults in California (McDonnell et al., 1999a; Greer et al., 1993). 

23 The Adventist Health and Smog (AHSMOG) study cohort of 3,914 (age 27 to 87 years, 

24 36% male) was drawn from nonsmoking, non-Hispanic white California Seventh Day 

25 Adventists, who were surveyed in 1977, 1987, and 1992. To be eligible, subjects had to 

26 have lived 10 or more years within 5 miles of their current residence in 1977. Residences 

27 from 1977 onward were followed and linked in time and space to interpolate 

concentrations of O3, PM10, SO4

2- 28 , SO2, and NO2. New asthma cases were defined as self29 reported doctor-diagnosed asthma at either the 1987 or 1992 follow-up questionnaire 

30 among those who had not reported having asthma upon enrollment in 1977. During the 

31 10-year follow-up (1977 to 1987), the incidence of new asthma was 2.1% for males and 

32 2.2% for females (Greer et al., 1993). Ozone concentration data were not provided. A 
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1 relative risk of 3.12 (95% CI: 1.16, 5.85) per 10-ppb increase in annual mean O3

2 (exposure metric not stated) was observed in males, compared to a relative risk of 0.94 

3 (95% CI: 0.65, 1.34) in females.

4 In the 15-year follow-up study (1977-1992), 3.2% of the eligible males and a slightly 

5 greater 4.3% of the eligible females developed adult asthma (McDonnell et al., 1999a). 

6 The mean 20-year average (1973-1992) for 8-h avg O3 (9 a.m. to 5 p.m.) was 46.5 ppb 

7 (SD 15.3). For males, the relative risk of developing asthma was 1.31 (95% CI: 1.01, 

8 1.71) per 10-ppb increase in 8-h avg O3. Once again, there was no evidence of a positive 

9 association between O3 and new-onset asthma in females (relative risk of 0.94 [95% CI: 

10 0.87, 1.02]). The lack of an association does not necessarily indicate no effect of O3 on 

11 the development of asthma among females. For example, differences between females 

12 and males in time-activity patterns may influence relative exposures to ambient O3. 

13 During summer 1992, the mean (SD) hours per week spent outdoors for male and female 

14 asthma cases were 13.8 (10.6) and 11.4 (10.9), respectively, indicating potential greater 

misclassification of exposure in females. None of the other pollutants (PM10, SO4

2- 15 , SO2, 

16 and NO2) were associated with development of asthma in either males or females. 

17 Adjusting for copollutants did not diminish the association between O3 and asthma 

18 incidence for males. In no case was the O3 coefficient reduced by more than 10% in the 

19 two-pollutant models compared to the model containing O3 alone. The consistency of the 

20 results in the two studies with different follow-up times, as well as the independent and 

21 robust association between annual mean O3 concentrations and asthma incidence, provide 

22 supportive evidence that long-term O3 exposure may be associated with the development 

23 of asthma in adult males. However, because the AHSMOG cohort was drawn from a 

24 narrow subject definition, the representativeness of this cohort to the general U.S. 

25 population may be limited.

26 In children, the relationship between long-term O3 exposure and new onset asthma has 

27 been extensively investigated in the CHS. In this cohort, evidence provides stronger 

28 support for long-term O3 exposure modifying the risk of new onset asthma associated 

29 with other potential risk factors than having a main effect on new onset asthma. Initiated 

30 in the early 1990s, the CHS was originally designed to examine whether long-term 

31 exposure to ambient pollutants was related to chronic respiratory outcomes in children in 

32 12 communities in southern California (Peters et al., 1999b; Peters et al., 1999a). New33 onset asthma was classified as having no prior history of asthma at study entry with 

34 subsequent report of physician-diagnosed asthma at follow-up with the date of onset 

35 assigned to be the midpoint of the interval between the interview date when asthma 

36 diagnosis was first reported and the previous interview date. In a cohort recruited during 

37 2002-2003 and followed for three years beginning in kindergarten or first grade, 

38 McConnell et al. (2010) reported a hazard ratio for new onset asthma of 0.76 (95% CI: 
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1 0.38, 1.54) comparing the communities with the highest (59.8 ppb) and lowest (29.5 ppb) 

2 annual average of 8-h avg (10 a.m.-6 p.m.) O3. With adjustment for school and residential 

3 modeled non-freeway traffic-related exposure, the estimated HR for O3 was 1.01 

4 (95% CI: 0.49, 2.11).

5 Similarly in a cohort recruited in 1993, asthma risk was not higher for residents of the six 

6 high-O3 communities versus residents of the six low-O3 communities (McConnell et al., 

7 2002). In this study, 3,535 initially nonasthmatic children (ages 9 to 16 years at 

8 enrollment) were followed for up to 5 years, during which 265 cases of new-onset asthma 

9 were identified. Communities were stratified by 4-year average 1-h max O3 levels, with 

10 six high-O3 communities (mean 75.4 ppb [SD 6.8]) and six low-O3 communities (mean 

11 50.1 ppb [SD 11.0]). Within the high-O3 communities, asthma risk was 3.3 (95% CI: 1.9, 

12 5.8) times greater for children who played three or more sports as compared with children 

13 who played no sports. None of the children who lived in high-O3 communities and played 

14 three or more sports had a family history of asthma. In models with individual sports 

15 entered as dummy variables, only tennis was significantly associated with asthma and 

16 only in the high O3 communities. This association was absent in the low-O3 communities 

17 (relative risk of 0.8 [95% CI: 0.4, 1.6]). The overall observed pattern of effects of sports 

18 participation on asthma risk was robust to adjustment for SES, history of allergy, family 

19 history of asthma, insurance, maternal smoking, and BMI.

20 Analyses aimed at distinguishing the effects of O3 from effects of other pollutants 

21 indicated that in communities with high O3 and low levels of other pollutants there was a 

22 4.2-fold (95% CI: 1.6, 10.7) increased risk of asthma in children playing three or more 

23 sports, compared to children who played no sports. The relative risk in children playing 

24 three or more sports was slightly lower (3.3 [95% CI: 1.6, 6.9]) in communities with a 

25 combination of high levels of O3 and other pollutants. Ozone concentrations were not 

26 strongly correlated with PM10, PM2.5, NO2, or inorganic acid vapors, and no associations 

27 with asthma were found for these other pollutants. These results provide additional 

28 support that the effects of physical activity on asthma are modified by long-term 

29 O3 exposure. Overall, the results from McConnell et al. (2002) suggest that playing sports 

30 may indicate greater outdoor activity when O3 levels are higher and an increased 

31 ventilation rate, which may lead to increased O3 exposure. It should be noted, however, 

32 that these findings were based on a small number of new asthma cases (n = 29 among 

33 children who played three or more sports) and were not based on a priori hypotheses. 

34 Recent studies from the CHS provide evidence for gene-environment interactions in 

35 effects on new-onset asthma by indicating that the lower risks associated with specific 

36 genetic variants are found in children who live in lower O3 communities (Islam et al., 

37 2009; Islam et al., 2008; Oryszczyn et al., 2007; Lee et al., 2004b; Gilliland et al., 2002). 
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1 Risk for new-onset asthma is related in part to genetic susceptibility, behavioral factors 

2 and environmental exposure (Gilliland et al., 1999). Gene-environment interactions in 

3 asthma have been well discussed in the literature (von Mutius, 2009; Holgate et al., 2007; 

4 Martinez, 2007a, b; Rahman et al., 2006; Hoffjan et al., 2005; Kleeberger and Peden, 

5 2005; Ober, 2005). Complex chronic diseases, such as asthma, are partially the result of a 

6 sequence of biochemical reactions involving exposures to various environmental agents 

7 metabolized by a number of different genes (Conti et al., 2003). Oxidative stress has been 

8 proposed to underlie these mechanistic hypotheses (Gilliland et al., 1999). Genetic 

9 variants may impact disease risk directly or modify disease risk by affecting internal dose 

10 of pollutants and other environmental agents and/or their reaction products or by altering 

11 cellular and molecular modes of action. Understanding the relation between genetic 

12 polymorphisms and environmental exposure can help identify high-risk subgroups in the 

13 population and provide better insight into pathway mechanisms for these complex 

14 diseases. 

15 CHS analyses have found that asthma risk is related to interactions between O3 and 

16 variants in genes for enzymes such as heme-oxygenase (HO-1), arginases (ARG1 and 2), 

17 and glutathione S transferase P1 (GSTP1) (Himes et al., 2009; Islam et al., 2008; Li et al., 

18 2008; Hanene et al., 2007; Ercan et al., 2006; Li et al., 2006a; Tamer et al., 2004; 

19 Gilliland et al., 2002). Biological plausibility for these findings is provided by evidence 

20 that these enzymes have antioxidant and/or anti-inflammatory activity and participate in 

21 well recognized modes of action in asthma pathogenesis. Further, several lines of 

22 evidence demonstrate that secondary oxidation products of O3 initiate the key modes of 

23 action that mediate downstream health effects (Section 5.3.2). For example, HO-1 has 

24 been found to respond rapidly to oxidants, have anti-inflammatory and anti-oxidant 

25 effects (Exner et al., 2004), relax airway smooth muscle, and be induced in airways 

26 during asthma (Carter et al., 2004). The GSTP1 Val/Val genotype has been associated 

27 with increased risk of having atopic asthma (Tamer et al., 2004). Gene-environment 

28 interactions are discussed in greater detail in Section 5.4.2.1. 

29 Islam et al. (2008) found that functional polymorphisms of the heme oxygenase-1 gene 

30 (HMOX-1, [(GT)n repeat]) influenced the risk of new-onset asthma, depending on 

31 ethnicity and long-term community O3 concentrations. Ozone-gene interactions were not 

32 found for variants in other antioxidant genes: catalase (CAT [-262C >T -844C >T0]) or 

33 and manganese superoxide dismutase (MNSOD, [Ala-9Val]). Analyses were restricted to 

34 children of Hispanic (n = 576) or non-Hispanic white ethnicity (n = 1,125) and were 

35 conducted with long-term pollutant levels averaged from 1994 to 2003. The effect of 

36 ambient air pollution on the relationship between genetic polymorphism and new-onset 

37 asthma was assessed using Cox proportional hazard regression models where the 

38 community specific average air pollution levels were fitted as a continuous variable 
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1 together with the appropriate interaction terms for genes and air pollutants and a random 

2 effect of community (Berhane et al., 2004).

3 Over the follow-up period, 160 new cases of asthma were diagnosed (Islam et al., 2008). 

4 For HMOX-1, the interaction (p = 0.003) indicated a greater protective effect of the 

5 S-allele (short, <23 (GT)n repeats) compared to the L-allele (long, >23 repeats) among 

6 non-Hispanic white children who lived in the low O3 community (nonparallelism 

7 presented in Figure 7-1). Among children residing in low-O3 communities, the hazard 

8 ratio (HR) of new onset asthma associated with the S-allele was 0.44 (95% CI: 0.23, 

9 0.83) compared to non-Hispanic white children who lived in low O3 communities and had 

10 no S-alleles. Biological plausibility for these results is provided by evidence that the 

11 S-allele variant of HMOX-1 is more readily induced than those with more numerous 

12 repeats. The S-allele was found to have a less protective effect in non-Hispanic white 

13 children who resided in high O3 communities (HR = 0.88; [95% CI: 0.33, 2.34] compared 

14 to non-Hispanic white children in low O3 communities with no S-allele). Because 

15 HMOX-1 variants were not associated with asthma risk in Hispanic children, effect 

16 modification by O3 was not investigated. No significant interactions were observed 

17 between PM10 or other pollutants and the HMOX-1 gene; quantitative results were not 

18 presented. Average O3 levels showed low correlation with the other monitored pollutants. 

19 The authors did not consider the lack of adjustment for multiple testing to be a concern in 

20 this analysis because the selection of the genes was based on a priori hypotheses defined 

21 by a well-studied biological pathway, in which oxidative stress serves as the link among 

22 O3 exposure, enzyme activity, and asthma.

23 Collectively, results from Islam et al. (2008) indicate that a variant in HMOX-1 that 

24 produces a more readily inducible enzyme is associated with lower risk of new-onset 

25 asthma in children who live in low O3 communities. Results were not presented for the 

26 main effects relating new-onset asthma to O3 exposure. However, they do indicate that 

27 that in environments of low ambient O3, enzymes with greater antioxidative activity may 

28 have the capacity to counter any temporary imbalance in an oxidant-antioxidant 

29 relationship. However, in the presence of high background O3, the protective effect may 

30 be attenuated because with higher exposure to oxidants, the antioxidant genes may be at 

31 their maximal level of inducibility, and variation in promoters no longer affects levels of 

32 expression. Supporting evidence is provided by Schroer et al. (2009), who found that 

33 infants with multiple environmental exposures were at increased risk of wheeze 

34 regardless of variant in GSTP1, which encodes a gene with antioxidant activity. 
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Note: An interaction p-value of 0.003 was obtained from the hierarchical two stage Cox proportional hazard model fitting the 

community specific O3 and controlling for random effect of the communities. The interaction indicates there is a greater protective 

effect of having a heme-oxygenase S-allele compared to having the L-allele among children living in communities with lower longterm ambient ozone concentrations. The HRs are off-set as opposed to overlapping in the figure to allow clearer presentation of the 

results. 

Source: Developed by EPA with data from Islam et al. (2008) (data used with permission of American Thoracic Society).

Figure 7-1 Interaction of heme-oxygenase genetic variants and O3 level on the 

Hazard Ratio (HR) of new-onset asthma in the 12 Children’s Health 

Study communities.

1 Expanding on the results of McConnell et al. (2002), Islam et al. (2009) provided 

2 evidence that variants in GSTM1 and GSTP1 may influence associations between 

3 outdoor exercise and new onset asthma. A primary conclusion that the authors (Islam et 

4 al., 2009) reported was that the GSTP1 Ile/Ile and GSTM1 null genotypes increased risk 

5 of new onset asthma during adolescence. The highest risk was found for participation in 

6 three or more team sports (compared to no sports) among children with GSTP1 Ile/Ile 

7 genotype living in high-O3 communities (HR: 6.15, [95% CI: 2.2, 7.4]). No three-way 

8 interaction was found for GSTM1. These results demonstrate the potential importance of 

9 a combination of genetic variability, O3 exposure, and outdoor activity on asthma risk. It 

10 is important to note that while some studies have found a modifying role of air pollution 

Interaction of Gene presence and Ozone Level on the 

Hazard Ratio of New Onset Asthma (P-value of 0.003)

0.94 0.88

1

0.44

0

0.5

1

1.5

2

2.5

3

Community Mean Ozone Level

HR

Children with no S-Allele

Children with S-Allele

 Low 

(38.4 ppb)

 High

(55.2 ppb)

(2.43) (2.34)

(0.36) (0.33)

(0.83)

(0.28)

Reference

(Confidence limits based on comparison with reference group)
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1 on the association between GSTP1 Ile/Ile and asthma in children (Lee et al., 2004b), 

2 others have found that the GSTP1 Val/Val variant to be associated with greater asthma 

3 prevalence and increase the risk of O3-associated respiratory morbidity (see discussion in 

4 Section 6.2.4.1).

5 The CHS also provided evidence of interactions between O3 exposure and variants in 

6 genes for arginase (Salam et al., 2009). Arginase catalyzes the conversion of L-arginine. 

7 Because L-arginine is a precursor of NO, higher arginase activity can limit production of 

8 NO and subsequent nitrosative stress. Epidemiologic evidence of associations of arginase 

9 variants with asthma are limited (Li et al., 2006a); however, asthmatic subjects have been 

10 found to have higher arginase activity than non-asthmatic subjects (Morris et al., 2004). 

11 The modifying effect of O3 and atopy on the association between ARG1 and ARG2 

12 haplotypes and asthma were evaluated using likelihood ratio tests with appropriate

13 interaction terms. Having more copies of the ARG1h4 haplotype (compared to having 

14 zero copies) was associated with lower odds of asthma, particularly among children with 

15 atopic asthma living in high O3 communities (OR: 0.12; [95% CI: 0.04, 0.43]). Having 

16 more copies of the ARG2h3 haplotype (compared to having zero copies) was associated 

17 with increased risk of childhood-onset asthma among children in both low and high O3

18 communities. The implications of findings are somewhat limited because the functional 

19 relevance of the ARG1 and ARG2 variants is not clear.

7.2.1.2 Prevalence of Asthma and Asthma Symptoms

20 Some cross-sectional studies reviewed in the 2006 O3 AQCD observed positive 

21 relationships between chronic exposure to O3 and prevalence of asthma and asthmatic 

22 symptoms in school children (Ramadour et al., 2000; Wang et al., 1999) while others 

23 (Kuo et al., 2002; Charpin et al., 1999) did not. Recent studies provide additional 

24 evidence.

25 In a cross-sectional nationwide study of 32,672 Taiwanese school children, Hwang et al. 

26 (2005) assessed the effects of air pollutants on the risk of asthma. The study population 

27 was recruited from elementary and middle schools within 1 km of air monitoring stations. 

28 The risk of asthma was related to O3 in the one-pollutant model. The addition of other 

29 pollutants (NOX, CO2, SO2, and PM10), in two-pollutant and three-pollutant models, 

30 increased the O3 risk estimates. The prevalence of childhood asthma was assessed in 

31 Portugal by contrasting the risk of asthma between a high O3 rural area and an area with 

32 low O3 levels (Sousa et al., 2011; Sousa et al., 2009; Sousa et al., 2008). The locations 

33 were selected to provide a difference in O3 levels without the confounding effects of 

34 other pollutants. Both evaluation for asthma symptoms and FEV1 suggested that O3
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1 increased asthma prevalence. Clark et al. (2010) investigated the effect of exposure to 

2 ambient air pollution in utero and during the first year of life on risk of subsequent 

3 incidence asthma diagnosis up to 3-4 years of age in a population-based nested case4 control study for all children born in southwestern British Columbia in 1999 and 2000 

5 (n = 37,401; including 3,482 [9.3%] with asthma). Air pollution exposure for each 

6 subject was estimated based on their residential address history using regulatory 

7 monitoring data, land use regression modeling, and proximity to stationary pollutant 

8 sources. Daily values from the three closest monitors within 50 km were used to calculate 

9 exposures. Traffic-related pollutants were associated with the highest risk. Ozone was 

10 inversely correlated with the primary traffic-related pollutants (r = -0.7 to -0.9). The low 

11 reliability of asthma diagnosis in infants makes this study difficult to interpret (Martinez 

12 et al., 1995). In a cross-sectional analysis, Akinbami et al. (2010) examined the 

13 association between chronic exposure to outdoor pollutants (12-month avg levels by 

14 county) and asthma outcomes in a national sample of children ages 3-17 years living in 

15 U.S. metropolitan areas (National Health Interview Survey, N = 34,073). A 5-ppb 

16 increase in estimated 8-h max O3 concentration (annual average) yielded a positive 

17 association for both currently having asthma and for having at least 1 asthma attack in the 

18 previous year, while the adjusted odds ratios for other pollutants were not statistically 

19 significant. Models in which pollutant value ranges were divided into quartiles produced 

20 comparable results. Multi-pollutant models (SO2 and PM) produced similar results. The 

21 median value for 12-month avg O3 levels was 39.5 ppb and the IQR was 35.9-43.7 ppb. 

22 The adjusted odds for current asthma for the highest quartile (49.9-59.5 ppb) of estimated 

23 O3 exposure was 1.56 (95% CI: 1.15, 2.10) with a positive concentration-response 

24 relationship apparent from the lowest quartile to the highest. Thus, this cross-sectional 

25 analysis and Hwang et al. (2005) provides further evidence relating O3 exposure and the 

26 risk of asthma.

27 Relationships between long-term exposure and respiratory symptoms in asthmatic 

28 children also were examined in the CHS. McConnell et al. (1999) examined the 

29 association between O3 levels and the prevalence of chronic lower respiratory tract 

30 symptoms in 3,676 cohort children with asthma. In this cross-sectional study, bronchitis, 

31 phlegm, and cough were not associated with annual mean 1-h max O3 concentrations in 

32 children with asthma or wheeze. All other pollutants examined (PM10, PM2.5, NO2, and 

33 gaseous acid) were associated with an increase in phlegm but not cough. The mean 

34 annual average 1-h max O3 concentration was 65.6 ppb (range 35.5 to 97.5) across the 

35 12 communities. In another CHS analysis, McConnell et al. (2003) evaluated 

36 relationships between air pollutants and bronchitic symptoms among 475 children with 

37 asthma. The mean 4-year average 8-h avg O3 (10 a.m.-6 p.m.) concentration was 

38 47.2 ppb (range 28.3 to 65.8) across the 12 communities. For a 10-ppb increase in 

39 8-h avg O3 averaged over 4 years, the between-community odds ratio was 0.90 (95% CI: 
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1 0.82, 1.00) whereas the within-community (i.e., difference between one- and four-year 

2 average) odds ratio was larger, i.e., 1.79 (95% CI: 1.00, 3.21). The authors commented 

3 that if the larger within-community effect estimates were correct, then other cross4 sectional (between-community) studies might have underestimated the true effect of air 

5 pollution on bronchitic symptoms in children. These differences might be attributable to 

6 confounding by poorly measured or unmeasured risk factors that vary between 

7 communities. Within community effects may more accurately represent risk associated 

8 with pollutant exposure because the analyses characterize health effects associated with 

9 changing pollutant concentrations within a community, thereby minimizing potential 

10 confounding by factors that are constant over time within a community. PM2.5, NO2, and 

11 organic carbon also were associated with bronchitic symptoms. In two-pollutant models, 

12 the within-community effect estimates for O3 were markedly reduced and no longer 

13 statistically significant in some cases. 

14 CHS also examined interactions between TNF-α 308 genotype and long-term O3

15 exposure in the occurrence of bronchitic symptoms among children with asthma (Lee et 

16 al., 2009b). Increased airway levels of the cytokine TNF-α has been related to 

17 inflammation, and the GG genotype has been linked to lower expression of TNF-α. 

18 Asthmatic children with the GG genotype had a lower prevalence of bronchitic symptoms 

19 compared with children carrying at least one A-allele (e.g., GA or AA genotype). Low20 versus high-O3 strata were defined as less than or greater than 50- ppb O3 avg. Asthmatic 

21 children with TNF-308 GG genotype had a significantly reduced risk of bronchitic 

22 symptoms with low-O3 exposure (OR: 0.53 [95% CI: 0.31, 0.91]). The risk was not 

23 reduced in children living in high-O3 communities (OR: 1.42 [95% CI: 0.75, 2.70]). The 

24 difference in genotypic effects between low- and high-O3 environments was statistically 

25 significant among asthmatics (P for interaction = 0.01), but not significant among non26 asthmatic children. Figure 7-2 presents adjusted O3 community-specific regression 

27 coefficients plotted against ambient O3 concentration, using weights proportional to the 

28 inverse variance. Investigators further reported no substantial differences in the effect of 

29 the GG genotype on bronchitic symptoms by long-term exposure to PM10, PM2.5, NO2, 

30 acid vapor, or second-hand smoke. 
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Note: Using indicator variables for each category of genotype and O3 exposure, investigators calculated effect estimates for TNF-α

GG genotype on the occurrence of bronchitic symptoms among children with asthma.

Source: Reprinted with permission of John Wiley & Sons, (Lee et al., 2009b).

Figure 7-2 Ozone modifies the effect of TNF GG genotype on bronchitic 

symptoms among children with asthma in the CHS.

1 Another CHS analyses reported interrelationships between variants in CAT and 

2 myeloperoxidase (MPO) genes, ambient pollutants, and respiratory-related school 

3 absences for 1,136 Hispanic and non-Hispanic white cohort children (Wenten et al., 

4 2009). A related study (Gilliland et al., 2001), found increased O3 exposure to be related 

5 greater school absenteeism due to respiratory illness but did not consider genetic variants. 

6 Wenten et al. (2009) hypothesized that variation in the level or function of antioxidant 

7 enzymes would modulate respiratory illness risk, especially under high levels of 

8 oxidative stress expected from high ambient O3 exposure. The joint effect of variants in 

9 these two genes (genetic epistasis) on respiratory illness was examined because the 

10 enzyme products operate on the same substrate within the same biological pathway. Risk 

11 of respiratory-related school absences was elevated for children with CAT GG plus MPO 

12 GA or AA genotypes (RR: 1.35 [95% CI: 1.03, 1.77] compared to GG for both genes) 

13 and reduced for children with CAT GA or AA plus MPO GA or AA (RR: 0.81 [95% CI: 

14 0.55, 1.19] compared to GG for both genes). Both CAT GG and MPO GA or AA 

15 genotypes produce a lower activity enzyme. In analyses that stratified communities into 

16 high and low O3 exposure groups by median levels (46.9 ppb), the protective effect of 

17 CAT GA or AA plus MPO GA or AA genotype was largely limited to children living in 
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1 communities with high ambient O3 levels (RR: 0.42 [95% CI: 0.20, 0.89]). The 

2 association of respiratory-illness absences with functional variants in CAT and MPO that 

3 differ by air pollution levels illustrates the need to consider genetic epistasis in assessing 

4 gene-environment interactions.

5 Collective evidence from CHS provides an important demonstration of gene-environment 

6 interactions. In the complex gene-environment setting a modifying effect might not be 

7 reflected in an exposure main effect. The simultaneous occurrence of main effect and 

8 interaction effect can occur. The study of gene-environment interactions helps to dissect 

9 disease mechanisms in humans by using information on susceptibility genes to focus on 

10 the biological pathways that are most relevant to that disease (Hunter, 2005). 

11 The French Epidemiology study on Genetics and Environment of Asthma (EGEA) 

12 investigated the relationship between ambient air pollution and asthma severity in a 

13 cohort in five French cities (Paris, Lyon, Marseille, Montpellier, and Grenoble) (Rage et 

14 al., 2009b). In this cross-sectional study, asthma severity over the past 12 months was 

15 assessed among 328 adult asthmatics using two methods: (1) a four-class severity score 

16 that integrated clinical events and type of treatment; and (2) a five-level asthma score 

17 based only on symptoms. Two measures of exposure were also assessed: (1) [first 

18 method]) closest monitor data from 1991 to 1995 where a total of 93% of the subjects 

19 lived within 10 km of a monitor, but where 70% of the O3 concentrations were 

20 back-extrapolated values; and (2) [second method]) a validated spatial model that used 

21 geostatistical interpolations and then assigned air pollutants to the geocoded residential 

22 addresses of all participants and individually assigned exposure to ambient air pollution 

23 estimates. Higher asthma severity scores were significantly related to both the 8-h avg O3

24 during April-September and the number of days with 8-h O3 averages above 55 ppb. Both 

25 exposure assessment methods and severity score methods resulted in very similar 

26 findings. Effect estimates of O3 were similar in three-pollutant models. No PM data were 

27 available. Since these estimates were not sensitive to the inclusion of ambient NO2 in the 

28 three-pollutant models, the authors viewed the findings not to be explained by particles 

29 which usually have substantial correlations between PM and NO2. Effect estimates for O3

30 in three-pollutant models including O3, SO2, and NO2 yielded OR for O3-days of 2.74 

31 (95% CI: 1.68, 4.48) per IQR days of 10-28 (+18) ppb. The effect estimates for SO2 and 

32 NO2 in the three-pollutant model were 1.33 (95% CI: 0.85, 2.11) and 0.94 (95% CI: 0.68, 

33 1.29) respectively. Taking into account duration of residence did not change the result. 

34 This study suggests that a higher asthma severity score is related to long-term O3

35 exposure.

36 An EGEA follow-up study (Jacquemin et al., In Press), examines the relationship 

37 between asthma and O3, NO2, and PM10. New aspects considered include: (1) 
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1 examination of three domains of asthma control (symptoms, exacerbations, and lung 

2 function); (2) levels of asthma control (controlled, partially controlled, and uncontrolled 

3 asthma); and (3) PM10 and multi-pollutant analysis. In this cross-sectional analysis, 

4 EGEA2 studied 481 adult subjects with current asthma from 2003 to 2007. The IQRs 

were 11 (41-52) μg/m3 for annual O3 and 13 (25-38) μg/m3 5 for summer (April6 September) O3. The association between asthma control and air pollutants was expressed 

7 by ORs (reported for one IQR of the pollutant), derived from multinomial logistic 

8 regression. For each factor, the simultaneous assessment of the risk for uncontrolled 

9 asthma and for partly controlled asthma was compared with controlled asthma using a 

10 composite of the three domains. In crude and adjusted models, O3-sum and PM10 were 

11 positively associated with partly controlled and uncontrolled asthma, with a clear gradient 

12 from controlled, partly controlled (OR = 1.53, 95% CI: 1.01, 2.33) and uncontrolled 

13 (OR = 2.14, 95% CI: 1.34, 3.43) (from the multinomial logistic regression). 

14 Separately, they used a composite asthma control classification that used the ordinal 

15 logistic regression for risk comparing controlled to partly controlled asthma and 

16 comparing partly controlled to uncontrolled asthma. For these two pollutants, the ORs

17 assessed using the ordinal logistic regression were significant (ORs were 1.69 (95% CI: 

18 1.22, 2.34) and 1.35 (95% CI: 1.13, 1.64) for O3-sum and PM10, respectively). For two 

19 pollutant models using the ordinal logistic regression, the adjusted ORs for O3-sum and 

20 PM10 included simultaneously in a unique model were 1.50 (95% CI: 1.07, 2.11) for O3-

21 sum and 1.28 (95% CI: 1.06, 1.55) for PM10, respectively. This result suggests that the 

22 effects of both pollutants are independent. 

23 The analysis of the associations between air pollution for all asthma subjects and each 

24 one of the three asthma control domains showed the following: (1) for lung function 

25 defined dichotomously as percent predicted FEV1 value <or >=80 (OR = 1.35, 95% CI: 

26 0.80, 2.28 for adjusted O3-sum); (2) for symptoms defined as asthma attacks or dyspnea 

27 or woken by asthma attack or shortness of breath in the past three months (OR = 1.59, 

28 95% CI: 1.10, 2.30 for adjusted O3-sum); and for exacerbations defined at least one 

29 hospitalizations or ER visits in the last year or oral corticosteroids in the past three 

30 months (OR = 1.58, 95% CI: 0.97, 2.59 for adjusted O3-sum). Since the estimates for 

31 both pollutants were more stable and significant when using the integrated measure of 

32 asthma control, this indicates that the results are not driven by one domain. These results 

33 support an effect of long-term exposure to O3 on asthma control in adulthood in subjects 

34 with pre-existing asthma.

35 Goss et al. (2004) investigated the effect of O3 on pulmonary exacerbations and lung 

36 function in individuals over the age of 6 years with cystic fibrosis (n = 11,484). The study 

37 included patients enrolled in the Cystic Fibrosis Foundation National Patient Registry, 
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1 which contains demographic and clinical data collected annually at accredited centers for 

2 cystic fibrosis. For 1999 through 2000, the annual mean O3 concentration, calculated 

3 from 1-h averages from 616 monitors in the U.S. EPA Aerometric Information Retrieval 

4 System (AIRS), was 51.0 ppb (SD 7.3). Exposure was assessed by linking air pollution 

5 values from AIRS with the patient’s home ZIP code. No clear association was found 

6 between annual mean O3 and lung function parameters. However, a 10 ppb increase in 

7 annual mean O3 was associated with a 10% (95% CI: 3, 17) increase in the odds of two or 

8 more pulmonary exacerbations. Significant excess odds of pulmonary exacerbations also 

9 were observed with increased annual mean PM10 and PM2.5 concentrations. The O3 effect 

10 was robust to adjustment for PM10 and PM2.5, 8% (95% CI: 1, 15) increase in odds of two 

11 or more pulmonary exacerbations per 10 ppb increase in annual mean O3.

7.2.2 Asthma Hospital Admissions and ED Visits

12 The studies on O3-related hospital discharges and emergency department (ED) visits for 

13 asthma and respiratory disease that were available in the 2006 O3 AQCD mainly looked 

14 at the daily time metric. Collectively the short-term O3 studies presented earlier in 

15 Section 6.2.7.5 indicate that there is evidence for increases in both hospital admissions 

16 and ED visits related to both all respiratory outcomes and asthma with stronger 

17 associations in the warm months. New studies evaluated long-term O3 exposure metrics, 

18 providing a new line of evidence that suggests a positive exposure-response relationship 

19 between first asthma hospital admission and long-term O3 exposure. 

20 An ecologic study (Moore et al., 2008) evaluated time trends in associations between 

21 declining warm-season O3 concentrations and hospitalization for asthma in children in 

22 California’s South Coast Air Basin who ranged in age from birth to 19 years. Quarterly 

23 average concentrations from 195 spatial grids, 10×10 km, were used. Ozone was the only 

24 pollutant associated with increased hospital admissions over the study period. A linear 

25 relation was observed for asthma hospital discharges (Moore et al., 2008). A matched 

26 case-control study (Karr et al., 2007) was conducted of infant bronchiolitis (ICD 9, code 

27 466.1) hospitalization and two measures of long-term pollutant exposure (the month prior 

28 to hospitalization and the lifetime average) for O3 in the South Coast Air Basin of 

29 southern California among 18,595 infants born between 1995 and 2000. Ozone was 

30 associated with reduced risk in the single-pollutant model, but this relation did not persist 

31 in multi-pollutant models (CO, NO2, and PM2.5).

32 In a cross-sectional study, Meng et al. (2010) examined associations between air 

33 pollution and asthma morbidity in the San Joaquin Valley in California by using the 2001 

34 California Health Interview Survey data from subjects ages 1 to 65+ who reported 
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1 physician-diagnosed asthma (n = 1,502). Subjects were assigned annual average 

2 concentrations for O3 based on residential ZIP code and the closet air monitoring station 

3 within 8 km but did not have data on duration of residence. Multi-pollutant models for O3

4 and PM did not differ substantially from single-pollutant estimates, indicating that 

5 pollutant multi-collinearity is not a problem in these analyses. The authors reported 

6 increased asthma-related ED visits or hospitalizations for O3 (OR = 1.49; [95% CI: 1.05, 

7 2.11] per 10 ppb) for all ages. Positive associations were obtained for symptoms, but 

8 95% confidence intervals included null values. Associations for symptoms for adults 

9 (ages 18 +) were observed (OR = 1.40; [95% CI: 1.02, 1.91] per 10 ppb).

10 Associations between air pollution and poorly controlled asthma among adults in 

11 Los Angeles and San Diego Counties were investigated using the California Health 

12 Interview Survey data collected between November 2000 and September 2001 (Meng et 

13 al., 2007). Each respondent was assigned an annual average concentration measured at 

14 the nearest station within 5 miles of the residential cross-street intersection. Poorly 

15 controlled asthma was defined as having daily or weekly asthma symptoms or at least one 

16 ED visit or hospitalization because of asthma during the past 12 months. This cross17 sectional study reports an OR of 3.34 (95% CI: 1.01, 11.09) for poorly controlled asthma 

18 when comparing those 65 years of age and older above the 90th percentile (28.7 ppb) 

19 level to those below that level. Co-pollutant (PM) analysis produced similar results.

20 Evidence associating long-term O3 exposure to first asthma hospital admission in a 

21 concentration-response relationship is provided in a retrospective cohort study (Lin et al., 

22 2008b). This study investigated the association between chronic exposure to O3 and 

23 childhood asthma admissions (defined as a principal diagnosis of ICD9, code 493) by 

24 following a birth cohort of 1,204,396 eligible births born in New York State during 

25 1995-1999 to first asthma admission or until 31 December 2000. There were 10,429 

26 (0.87%) children admitted to the hospital for asthma between 1 and 6 years of age. The 

27 asthma hospitalization rate in New York State in 1993 was 2.87 per 1,000 (Lin et al., 

28 1999). Three annual indicators (all 8-h max from 10:00 a.m. to 6:00 p.m.) were used to 

29 define chronic O3 exposure: (1) mean concentration during the follow-up period 

30 (41.06 ppb); (2) mean concentration during the O3 season (50.62 ppb); and (3) proportion 

31 of follow-up days with O3 levels >70 ppb. In this study the authors aimed to predict the 

32 risk of having asthma admissions in a birth cohort, but the time to the first admission in 

33 children that is usually analyzed in survival models was not their primary interest. The 

34 effects of co-pollutants were assessed and controlled for using the Air Quality Index 

35 (AQI). Interaction terms were used to assess potential effect modifications. A positive 

36 association between chronic exposure to O3 and childhood asthma hospital admissions 

37 was observed indicating that children exposed to high O3 levels over time are more likely 

38 to develop asthma severe enough to be admitted to the hospital. The various factors were 
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1 examined and differences were found for younger children (1-2 years), poor 

2 neighborhoods, Medicaid/self-paid births, geographic region and others. As shown in 

3 Figure 7-3, positive concentration-response relationships were observed. Asthma 

4 admissions were significantly associated with increased O3 levels for all chronic exposure 

5 indicators (ORs, 1.16-1.68). When estimating the O3 effect using the exceedance 

6 proportion, an increase was observed (OR = 1.68; [95% CI: 1.64, 1.73]) in hospital 

7 admissions with an IQR (2.51%) increase in O3. A proportional hazards model for the 

8 New York City data was run as a sensitivity analysis and it yielded similar results 

9 between asthma admissions and chronic exposure to O3 (Cox model: HR = 1.14, 

10 [95% CI: 1.124, 1.155] is similar to logistic model results: OR = 1.16 [95% CI: 1.15, 

11 1.17]) (Lin, 2010). Thus, this study provides evidence associating long-term O3 exposure 

12 to first asthma hospital admission in a concentration-response relationship. 

13 In considering relationships between long-term pollutant exposure and chronic disease 

14 health endpoints, Künzli (2012) offers two hypotheses relevant to research on air 

15 pollution and chronic disease where chronic pathologies are found with acute expressions 

16 of the chronic disease: “H1: Exposure provides a basis for the development of the 

17 underlying chronic pathology, which increases the pool of people with chronic conditions 

18 prone to exacerbations; H2: Exposure triggers an acute event (or a state of frailty that 

19 results in an event with a delay of a few days or weeks) among those with the disease.” 

20 Künzli (2012) states if associations of pollution with events are much larger in the long21 term studies, it provides some indirect evidence in support of H1. If air pollution 

22 increases the pool of subjects with the chronic pathology (H1), more acute events are 

23 expected to be seen for higher exposures since events due to various causes are part of the 

24 chronic disease pathway.

25 Künzli (2012) makes such a comparison noting larger associations with long-term NO2

26 exposures for adult asthma hospital admissions (Andersen et al., 2012) as compared to 

27 short-term NO2 exposures for asthma hospital admissions (Peel et al., 2005). In a further 

28 example, Pope (2007) makes similar conclusions comparing long-term PM mortality 

29 study results to short-term PM mortality studies. The results of Lin et al. (2008b) for first 

30 asthma hospital admission, presented below, show effect estimates that are larger than 

31 those reported in a study of asthma hospital admissions in New York State by Silverman 

32 and Ito (2010), discussed in Chapter 6 (both studies are for young children). This 

33 provides some support for the hypothesis that O3 exposure may not only have triggered 

34 the events but also increased the pool of asthmatics. 
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Note: Adjusted for child’s sex, age, birth weight, and gestational age; maternal race, ethnicity, age, education, insurance, and 

smoking status during pregnancy; and regional poverty level and temperature. ORs by low, medium, and high exposure are shown 

for New York City (NYC: low [37.3 ppb], medium [37.3-38.11 ppb], high [38.11+ ppb] and other New York State regions (Other NYS 

regions: low [42.58 ppb], medium [42.58-45.06 ppb], high [45.06+ ppb]) for first asthma hospital admission.

Source: Lin (2010); Lin et al. (2008b)

Figure 7-3 Ozone-asthma concentration-response relationship using the mean 

concentration during the entire follow-up period for first asthma 

hospital admission.

7.2.3 Pulmonary Structure and Function

7.2.3.1 Pulmonary Structure and Function: Evidence from 

Epidemiology Studies

1 The definitive 8-year follow-up analysis of the first cohort of the CHS, which is 

2 discussed in Section 7.2 (Gauderman et al., 2004), provided little evidence that long-term 

3 exposure to ambient O3 was associated with significant deficits in the growth rate of lung 

4 function in children. A later CHS study (Islam et al., 2007) examined relationships 

5 between air pollution, lung function, and new onset asthma and reported no substantial 

6 differences in the effect of O3 on lung function. Ozone concentrations from the least to 

7 most polluted communities (mean annual average of 8-h avg O3) ranged from 30 to 

8 65 ppb, as compared to the ranges observed for the other pollutants, which had 4-fold- to 

9 8-fold differences in concentrations. In a more recent CHS study, Breton et al. (2011)

10 hypothesized that genetic variation in genes on the glutathione metabolic pathway may 
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1 influence the association between ambient air pollutant exposures and lung function 

2 growth in children. They investigated whether genetic variation in glutathione genes 

3 GSS, GSR, GCLC, and GCLM was associated with lung function growth in healthy 

4 children using data collected on 2,106 children over an 8-year time-period as part of the 

5 Children’s Health Study. Breton et al. (2011) found that variation in the GSS locus was 

6 associated with differences in risk of children for lung function growth deficits associated 

7 with NO2, PM10, PM2.5, elemental carbon, organic carbon, and O3. The negative effects of 

8 air pollutants were largely observed within participants who had a particular GSS 

9 haplotype. The effects ranged from -124.2 to -149.1 mL for FEV1, -92.9 to -126.7 mL for 

10 FVC and -193.9 to -277.9 mL/sec for MMEF for all pollutants except O3, for which some 

11 positive associations were reported: 25.9 mL for FEV1; 0.1 mL for FVC, and 

12 166.5 mL/sec for MMEF. Ozone was associated with larger decreases in lung function in 

13 children without this haplotype, when compared to the other pollutants with values of 

14 -76.6 mL for FEV1, -17.2 mL for FVC, and -200.3 mL/sec for MMEF, but only the 

15 association with MMEF was statistically significant.

16 As discussed in the 2006 O3 AQCD, a study of freshman students at the University of 

17 California, Berkeley reported that lifetime exposure to O3 was associated with decreased 

18 measures of small airways (<2 mm) function (FEF75 and FEF25-75) (Tager et al., 2005). 

19 There was an interaction with the FEF25-75/FVC ratio, a measure of intrinsic airway size. 

20 Subjects with a large ratio (indicating an increased airway size relative to their lung 

21 volume) were less likely to have decreases in FEF75 and FEF25-75 for a given estimated 

22 lifetime exposure to O3. Kinney and Lippmann (2000) examined 72 nonsmoking adults 

23 (mean age 20 years) from the second-year class of students at the U.S. Military Academy 

24 in West Point, NY, and reported results that appear to be consistent with a decline in lung 

25 function that may in part be due to O3 exposures over a period of several summer months. 

26 Ihorst et al. (2004) examined 2,153 children with a median age of 7.6 years and reported 

27 pulmonary function results which indicated that significantly lower FVC and FEV1

28 increases were associated with higher O3 exposures over the medium-term of several 

29 summer months, but not over several months in the winter. Semi-annual mean O3

30 concentrations ranged from 22 to 54 ppb during the summer months and 4 to 36 ppb 

31 during the winter months. Further, over the longer-term 3.5-year period Ihorst et al. 

32 (2004) found that higher mean summer months O3 levels were not associated with growth 

33 rates in lung function and for FVC and FEV1, in contrast to the significant medium-term 

34 effects. Frischer et al. (1999) found that higher O3 over one summer season, one winter 

35 season, and greater increases from one summer to the next over a three-year period were 

36 associated with smaller increases in lung function growth, indicating both medium and 

37 longer-term effects.
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1 (Mortimer et al., 2008a, b) examined the association of prenatal and lifetime exposures to 

2 air pollutants with pulmonary function and allergen sensitization in a subset of asthmatic 

3 children (ages 6-11) included in the Fresno Asthmatic Children’s Environment Study 

4 (FACES). Monthly means of pollutant levels for the years 1989-2000 were created and 

5 averaged separately across several important developmental time-periods, including: the 

6 entire pregnancy, each trimester, the first 3 years of life, the first 6 years of life, and the 

7 entire lifetime. In the first analysis (Mortimer et al., 2008a), negative effects on 

8 pulmonary function were found for exposure to PM10, NO2, and CO during key neonatal 

9 and early life developmental periods. The authors did not find a negative effect of 

10 exposure to O3 within this cohort. In the second analysis (Mortimer et al., 2008b), 

11 sensitization to at least one allergen was associated, in general, with higher levels of CO 

12 and PM10 during the entire pregnancy and second trimester, and higher PM10 during the 

13 first 2 years of life. Lower exposure to O3 during the entire pregnancy or second trimester 

14 was associated with an increased risk of allergen sensitization. Although the pollutant 

15 metrics across time periods were correlated, the strongest associations with the outcomes 

16 were observed for prenatal exposures. Though it may be difficult to disentangle the effect 

17 of prenatal and postnatal exposures, the models from this group of studies suggest that 

18 each time period of exposure may contribute independently to different dimensions of 

19 school-aged children’s pulmonary function. For 4 of the 8 pulmonary-function measures 

20 (FVC, FEV1, PEF, FEF25-75), prenatal exposures were more influential on pulmonary 

21 function than early-lifetime metrics, while, in contrast, the ratio of measures (FEV1/FVC 

22 and FEF25-75/FVC) were most influenced by postnatal exposures. When lifetime metrics 

23 were considered alone, or in combination with the prenatal metrics, the lifetime measures 

24 were not associated with any of the outcomes. This suggests that the timing of the O3

25 exposure may be more important than the overall dose, and prenatal exposures are not 

26 just markers for lifetime or current exposures.

27 Latzin et al. (2009) examined whether prenatal exposure to air pollution was associated 

28 with lung function changes in the newborn. Tidal breathing, lung volume, ventilation 

29 inhomogeneity and eNO were measured in 241 unsedated, sleeping neonates 

30 (age = 5 weeks). Consistent with the previous studies, no association was found for 

31 prenatal exposure to O3 and lung function.

32 In a cross-sectional study of adults, Qian et al. (2005) examined the association of long33 term exposure to O3 and PM10 with pulmonary function from data of 10,240 middle-aged 

34 subjects who participated in the Atherosclerosis Risk in Communities (ARIC) study in 

35 four U.S. communities. A surrogate for long-term O3 exposure from daily data was 

36 determined at the individual level. Ozone was significantly and negatively associated 

37 with measures of pulmonary function. 
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1 To determine the extent to which long-term exposure to outdoor air pollution accelerates 

2 adult decline in lung function, Forbes et al. (2009b) studied the association between 

3 chronic exposure to outdoor air pollution and lung function in approximately 42,000 

4 adults aged 16 and older who were representatively sampled cross-sectionally from 

5 participants in the Health Survey for England (1995, 1996, 1997, and 2001). FEV1 was 

6 not associated with O3 concentrations. In contrast to the results for PM10, NO2, and SO2;

7 combining the results of all the survey years showed that a 5-ppb difference in O3 was 

8 counter-intuitively associated with a higher FEV1 by 22 mL.

9 In a prospective cohort study consisting of school-age, non-asthmatic children in 

10 Mexico City (n = 3,170) who were 8 years of age at the beginning of the study, Rojas11 Martinez et al. (2007) evaluated the association between long-term exposure to O3, PM10

12 and NO2 and lung function growth every 6 months from April 1996 through May 1999. 

13 Exposure data were provided by 10 air quality monitor stations located within 2 km of 

14 each child’s school. Over the study period, 8-h O3 concentrations ranged from 60 ppb 

15 (SD, ± 25) in the northeast area of Mexico City to 90 ppb (SD, ± 34) in the southwest, 

16 with an overall mean of 69.8 ppb. In multi-pollutant models, an IQR increase in mean O3

17 concentration of 11.3 ppb was associated with an annual deficit in FEV1 of 12 mL in girls 

18 and 4 mL in boys. Single-pollutant models showed an association between ambient 

19 pollutants (O3, PM10, and NO2) and deficits in lung function growth. While the estimates 

20 from co-pollutant models were not substantially different than single pollutant models, 

21 independent effects for pollutants could not be estimated accurately because the traffic22 related pollutants were correlated. To reduce exposure misclassification, 

23 microenvironmental and personal exposure assessments were conducted in a randomly 

24 selected subsample of 60 children using passive O3 samplers. Personal O3 concentrations 

25 were correlated (p <0.05) with the measurements obtained from the fixed-site air 

26 monitoring stations. 

27 In the 2006 O3 AQCD, few studies had investigated the effect of chronic O3 exposure on 

28 pulmonary function. The strongest evidence was for medium-term effects of extended O3

29 exposures over several summer months on lung function (FEV1) in children, i.e., reduced 

30 lung function growth being associated with higher ambient O3 levels. Longer-term 

31 studies (annual), investigating the association of chronic O3 exposure on lung function 

32 (FEV1) such as the definitive 8-year follow-up analysis of the first cohort (Gauderman et 

33 al., 2004) provides little evidence that long-term exposure to ambient O3 at current levels 

34 is associated with significant deficits in the growth rate of lung function in children. 

35 Analyses indicated that there was no evidence that either 8-h avg O3 (10 a.m. to 6 p.m.) 

36 or 24-h avg O3 was associated with any measure of lung function growth over a 4-year 

37 (age 10 to 14 years; (Gauderman et al., 2000)) or 8-year (age 10 to 18 years; (Gauderman 

38 et al., 2004)) period. However, most of the other pollutants examined (including PM2.5, 
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1 NO2, acid vapor, and elemental carbon) were found to be significantly associated with 

2 reduced growth in lung function. In addition, there was only about a 2- to 2.5-fold 

3 difference in O3 concentrations from the least to most polluted communities (mean 

4 annual average of 8-h avg O3 ranged from 30 to 65 ppb), versus the ranges observed for 

5 the other pollutants (which had 4- to 8-fold differences in concentrations). 

6 Short-term O3 exposure studies presented in Section 6.2.1.2 provide a cumulative body of 

7 epidemiologic evidence that strongly supports associations between ambient O3 exposure 

8 and decrements in lung function among children. For new studies of long-term O3

9 exposure relationship to pulmonary function, one study, where O3 and other pollutant 

10 levels were higher (90 ppb at high end of the range) than those in the CHS, observes a 

11 relationship between O3 concentration and pulmonary function declines in school-aged 

12 children. Two studies of adult cohorts provide mixed results where long- term exposures 

13 were at the high end of the range with levels of 49.5 ppb in one study and 27 ppb IQR in 

14 the other. Toxicological studies examining monkeys have provided data for airway 

15 resistance in an asthma model but this is difficult to compare to FEV1 results. Thus there 

16 is little new evidence to build upon the very limited studies of pulmonary function 

17 (FEV1) from the 2006 O3 AQCD. 

7.2.3.2 Pulmonary Structure and Function: Evidence from 

Toxicological Studies and Nonhuman Primate Asthma 

Models

18 Long-term studies in animals allow for greater insight into the potential effects of 

19 prolonged exposure to O3, that may not be easily measured in humans, such as structural 

20 changes in the respiratory tract. As reviewed in the 1996 and 2006 O3 AQCDs and 

21 Chapter 5 of this ISA, there are both qualitative and quantitative uncertainties in the 

22 extrapolation of data generated by rodent toxicology studies to the understanding of 

23 health effects in humans. Despite these uncertainties, epidemiologic studies observing 

24 functional changes in humans can attain biological plausibility, in conjunction with long25 term toxicological studies, particularly O3-inhalation studies performed in non-human 

26 primates whose respiratory system most closely resembles that of the human. An 

27 important series of studies have used nonhuman primates to examine the effect of O3

28 alone or in combination with an inhaled allergen, house dust mite antigen, on 

29 morphology and lung function. These animals exhibit the hallmarks of allergic asthma 

30 defined for humans, including: a positive skin test for HDMA with elevated levels of IgE 

31 in serum and IgE-positive cells within the tracheobronchial airway walls; impaired 

32 airflow which is reversible by treatment with aerosolized albuterol; increased abundance 

33 of immune cells, especially eosinophils, in airway exudates and bronchial lavage; and 
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1 development of nonspecific airway responsiveness (NHLBI, 2007). Hyde et al. (2006)

2 compared asthma models of rodents (mice) and the nonhuman primate model to 

3 responses in humans and concluded that the unique responses to inhaled allergen shown 

4 in the rhesus monkeys make it the most appropriate animal model of human asthma. 

5 These studies and others have demonstrated changes in pulmonary function and airway 

6 morphology in adult and infant nonhuman primates repeatedly exposed to 

7 environmentally relevant concentrations of O3 (Joad et al., 2008; Carey et al., 2007; 

8 Plopper et al., 2007; Fanucchi et al., 2006; Joad et al., 2006; Evans et al., 2004; Larson et 

9 al., 2004; Tran et al., 2004; Evans et al., 2003; Schelegle et al., 2003; Fanucchi et al., 

10 2000; Hyde et al., 1989; Harkema et al., 1987a; Harkema et al., 1987b; Fujinaka et al., 

11 1985). Many of the observations found in adult monkeys have also been noted in infant 

12 rhesus monkeys, although a direct comparison of the degree of effects between adult and 

13 infant monkeys has not been reported. The findings of these nonhuman primate studies 

14 have also been observed in rodent studies discussed at the end of this section and 

15 included in Table 7-1.

16 The initial observations in adult nonhuman primates have been expanded in a series of 

17 experiments using infant rhesus monkeys repeatedly exposed to 0.5 ppm O3 starting at 

1 month of age1 18 (Plopper et al., 2007). The purpose of these studies, designed by Plopper 

19 and colleagues, was to determine if a cyclic regimen of O3 inhalation would amplify the 

20 allergic responses and structural remodeling associated with allergic sensitization and 

21 inhalation in the infant rhesus monkey. In terms of pulmonary function changes, after 

22 several episodic exposures of infant monkeys to O3, they observed a significant increase 

23 in the baseline airway resistance, which was accompanied by a small increase in airway 

24 responsiveness to inhaled histamine (Schelegle et al., 2003), although neither 

25 measurement was statistically different from filtered air control values. Exposure of 

26 animals to inhaled house dust mite antigen alone also produced small but not statistically 

27 significant changes in baseline airway resistance and airway responsiveness, whereas the 

28 combined exposure to both (O3 + antigen) produced statistically significant and greater 

29 than additive changes in both functional measurements. This nonhuman primate evidence 

30 of an O3-induced change in airway resistance and responsiveness supports the biologic 

31 plausibility of long-term exposure to O3 contributing to the effects of asthma in children. 

32 To understand which conducting airways and inflammatory mechanisms are involved in 

33 O3-induced airway hyperresponsiveness in the infant rhesus monkey, a follow-up study 

34 examined airway responsiveness ex vivo in lung slices (Joad et al., 2006). Using video 

35 microscopy to morphometrically evaluate the response of bronchi and respiratory 

 1 Schelegle et al. (2003) used a two-by-two block design. Twenty-four infant rhesus monkeys (30 days old) were exposed to 11 

episodes (total of 6-months exposure period) of filtered air (FA), house dust mite allergen (HDMA), O3 (5 days each followed by 9 

days of FA). Ozone was delivered for 8h/day at 0.5 ppm. Twelve of the monkeys (HDMA, and HDMA + O3 groups) were sensitized 

to house dust mite allergen (HDMA, confirmed by skin testing). To evaluate the potential for recovery, the 5 months of exposure 

were followed by another 6 months in FA until the monkeys were reevaluated at 12 months of age.
Draft – Do Not Cite or Quote 7-24 June 2012

1 bronchioles to methacholine, (a bronchoconstricting agent commonly used to evaluate 

2 airway responsiveness in asthmatics), the investigators observed differential effects for 

3 the two airway sizes. While episodic exposure to O3 alone (0.5 ppm) had little effect on 

4 ex vivo airway responsiveness in bronchi and respiratory bronchioles, exposure to dust 

5 mite antigen alone produced airway hyperresponsiveness in the large bronchi, whereas 

6 O3 + antigen produced significant increases in airway hyperresponsiveness only in the 

7 respiratory bronchioles. These results suggest that ozone’s effect on airway 

8 responsiveness occurs predominantly in the smaller bronchioles, where dosimetric 

9 models indicate the dose would be higher.

10 The functional changes in the conducting airways of infant rhesus monkeys exposed to 

11 either O3 alone or O3 + antigen were accompanied by a number of cellular and 

12 morphological changes, including a significant 4-fold increase in eosinophils, (a cell type 

13 important in allergic asthma), in the bronchoalveolar lavage of infant monkeys exposed 

14 to O3 alone. Thus, these studies demonstrate both functional and cellular changes in the 

15 lung of infant monkeys after cyclic exposure to 0.5 ppm O3. This concentration, provides 

16 relevant information to understanding the potentially damaging effects of ambient O3

17 exposure on the respiratory tract of humans. No concentration-response data, however, 

18 are available from these nonhuman primate studies.

19 In addition to these functional and cellular changes, significant structural changes in the 

20 respiratory tract have been observed in infant rhesus monkeys exposed to O3. During 

21 normal respiratory tract development, conducting airways increase in diameter and length 

22 in the infant rhesus monkey. Exposure to O3 alone (5 days of 0.5 ppm O3 at 8 h/day, 

23 followed by 9 days of filtered air exposures for 11 cycles), however, markedly affected 

24 the growth pattern of distal conducting airways (Fanucchi et al., 2006). Whereas the first 

25 alveolar outpocketing occurred at airway generation 13 or 14 in filtered air-control infant 

26 monkeys, the most proximal alveolarized airways occurred at an average of 10 airway 

27 generations in O3-exposed monkeys. Similarly, the diameter and length of the terminal 

28 and respiratory bronchioles were significantly decreased in O3-exposed monkeys. 

29 Importantly, the O3-induced structural pathway changes persisted after recovery in 

30 filtered air for 6 months after cessation of the O3 exposures. These structural effects were 

31 accompanied by significant increases in mucus goblet cell mass, alterations in smooth 

32 muscle orientation in the respiratory bronchioles, epithelial nerve fiber distribution, and 

33 basement membrane zone morphometry. These latter effects are noteworthy because of 

34 their potential contribution to airway obstruction and airway hyperresponsiveness which 

35 are central features of asthma.

36 Because many cellular and biochemical factors are known to contribute to allergic 

37 asthma, the effect of exposure to O3 alone or O3 + antigen on immune system parameters 
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1 was also examined in infant rhesus monkeys. Mast cells, which contribute to asthma via 

2 the release of potent proteases, were elevated in animals exposed to antigen alone but O3

3 alone had little effect on mast cell numbers and the response of animals exposed to O3 + 

4 antigen was not different from that of animals exposed to antigen alone; thus suggesting 

5 that mast cells played little role in the interaction between O3 and antigen in this model of 

6 allergic asthma (VanWinkle et al., 2010). Increases in CD4+ and CD8+ lymphocytes 

7 were observed at 6 months of age in the blood and bronchoalveolar lavage fluid of infant 

8 rhesus monkeys exposed to O3 + antigen but not in monkeys exposed to either agent 

9 alone (Miller et al., 2009). Activated lymphocytes (i.e., CD25+ cells) were 

10 morphometrically evaluated in the airway mucosa and significantly increased in infant 

11 monkeys exposed to antigen alone or O3 + antigen. Although O3 alone had no effect on 

12 CD25+ cells, it did alter the anatomic distribution of CD25+ cells within the airways. 

13 Ozone had only a small effect on these sets of immune cells and did not produce a strong 

14 interaction with an inhaled allergen in this nonhuman primate model.

15 In addition to alterations in the immune system, nervous system interactions with 

16 epithelial cells are thought to play a contributing role to airway hyperresponsiveness. A 

17 critical aspect of postnatal lung development is the laying of nerve axons with specific 

18 connections serving to maintain lung homeostasis. Aberrant innervation patterns may 

19 underlie allergic airways disease pathology and long-term decrements in airway function. 

20 As noted in the 2006 O3 AQCD, exposure of infant rhesus monkeys altered the normal 

21 development of neural innervation in the epithelium of the conducting airways (Larson et 

22 al., 2004). Significant mean reductions in nerve fiber density were observed in the 

23 midlevel airways of animals exposed to O3 alone (49% reduction), and O3 + antigen (55% 

24 reduction). Moreover, the morphology of nerve bundles was altered. The persistence of 

25 these effects was examined after a 6-month recovery period, and although nerve 

26 distribution remained atypical, there was a dramatic increase in airway nerve density 

27 (hyperinnervation) (Kajekar et al., 2007). Thus, in addition to structural, immune, and 

28 inflammatory effects, exposure to O3 produces alterations in airway innervation which 

29 may contribute to O3-induced exacerbation of asthma. Evaluation of the pathobiology of 

30 airway remodeling in growing lungs of neonates using an animal model where exposure 

31 to allergen generates reactive airway disease with all the hallmarks of asthma in humans 

32 illustrates that exposure to O3 and allergen early in life produces a large number of 

33 disruptions of fundamental growth and differentiation processes. 

34 A number of studies in both nonhuman primates and rodents demonstrate that O3

35 exposure can increase collagen synthesis and deposition, inducing fibrotic-like changes in 

36 the lung (Last et al., 1994; Chang et al., 1992; Moffatt et al., 1987; Reiser et al., 1987; 

37 Last et al., 1984). Increased collagen content is often associated with elevated abnormal 

38 cross links that appear to be irreversible (Reiser et al., 1987). Generally changes in 
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1 collagen content have been observed in rats exposed to 0.5 ppm O3 or higher, although 

2 extracellular matrix thickening has been observed in the lungs of rats exposed to an urban 

3 pattern of O3 with daily peaks of 0.25 ppm for 38 weeks (Chang et al., 1992; Chang et al., 

4 1991). A more recent study using an urban pattern of exposure to 0.5 ppm O3

5 demonstrated that O3-induced collagen deposition in mice is dependent on the activity of 

6 TGF-β (Katre et al., 2011). Sex differences have been observed with respect to increased 

7 centriacinar collagen deposition and crosslinking, which was observed in female but not 

8 male rats exposed to 0.5 and 1.0 ppm O3 for 20 months (Last et al., 1994). Few other 

9 long-term exposure morphological studies have presented sex differences and most only 

10 evaluated males. 

11 As described in the 1996 and 2006 O3 AQCDs, perhaps the largest chronic O3 study was 

12 an NIEHS-NTP/HEI funded rodent study conducted by multiple investigators studying a 

13 number of different respiratory tract endpoints (Catalano et al., 1995b). Rats were 

14 exposed to 0.12, 0.5, or 1.0 ppm O3 for 6 h/day and 5 d/week for 20 months. The most 

15 prominent changes were observed in the nasal cavity where a large fraction of O3 is 

16 absorbed. Alterations in nasal function (increased mucous flow) and structure (goblet cell 

17 metaplasia) were observed at 0.5 and 1.0 ppm but not 0.12 ppm O3. In the lung, the 

18 centriacinar region (CAR) was the anatomical site most affected by O3. The epithelial cell 

19 lining was changed to resemble that seen in respiratory bronchioles and the interstitial 

20 volume was increased. Biochemical analyses demonstrated increased collagen and 

21 glycoaminoglycans, an observation that supported the structural changes. As in the nose, 

22 these changes were observed only at the two highest exposure concentrations. 

23 Importantly, despite these morphologic and biochemical changes after 20 months of 

24 exposure, detailed pulmonary function testing revealed little to no measurable change in 

25 function. Thus, minor respiratory tract changes were observed after chronic exposure to 

26 O3 up to 1.0 ppm in the F344 rat model. 

27 It is unclear what the long-term impact of O3-induced structural changes may be. 

28 Simulated seasonal (episodic) exposure studies suggest that such exposures might have 

29 cumulative impacts, and a number of studies indicate that structural changes in the 

30 respiratory system are persistent or irreversible. For example, O3-induced hyperplasia 

31 was still evident in the nasal epithelia of rats 13 weeks after recovery from 0.5 ppm O3

32 exposure (Harkema et al., 1999). In a study of episodic exposure to 0.25 ppm O3, Chang 

33 et al. (1992) observed no reversal of basement membrane thickening in rat lungs up to 17 

34 weeks post-exposure. Thickening of the sub-basement membrane is one of the persistent 

35 structural features observed in human asthmatics (NHLBI, 2007). Episodic exposure 

36 (0.25 ppm O3, every other month) of young monkeys induced equivalent morphological 

37 changes compared to continuously exposed animals, even though they were exposed for 

38 half the time and evaluation occurred a month after exposure ceased as opposed to 
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1 immediately (Tyler et al., 1988). Notably, episodic O3 exposure increased total lung 

2 collagen content, chest wall compliance, and inspiratory capacity, suggesting a delay in 

3 lung maturation in episodically-exposed animals. These changes were in contrast to the 

4 continuously exposed group, which did not differ from the air exposed group in these 

5 particular parameters but did exhibit greater bronchiolitis than the episodically exposed 

6 animals. In a study by Harkema and colleagues (Harkema et al., 1993, 1987b), monkeys 

7 (both males and females) were acutely exposed for 8 h/day to 0.15 ppm O3 (6 days) or 

8 chronically to 0.15 ppm or 0.3 ppm O3 (90 days). For most endpoints in the nasal cavity, 

9 the observed morphologic changes and inflammation were greater in the monkeys 

10 exposed for 6 days compared to 90 days, whereas in the respiratory bronchioles of the 

11 same animals, there were no significant time or concentration dependent differences 

12 (increased epithelial thickness and proportion of cuboidal cells) between the 6 and 90 day 

13 exposure groups.

14 Stokinger (1962) reported that chronic bronchitis, bronchiolitis, and emphysematous and 

15 fibrotic changes develop in the lung tissues of mice, rats, hamsters, and guinea pigs 

16 exposed 6 h/day, 5 days/week for 14.5 months to a concentration slightly above 1 ppm 

17 O3. Rats continuously exposed for 3 to 5 months to 0.8 ppm O3 develop a disease that 

18 resembles emphysema, and they finally die of respiratory failure (Stephens et al., 1976). 

19 Ozone results in a greater response of fibroblasts in the lesion, thickening of the alveolar 

20 septae, and an increase in number of alveolar macrophages in the proximal alveoli. 
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Table 7-1 Respiratory effects in nonhuman primates and rodents resulting 

from long-term ozone exposure

Study Model O3 (ppm) Exposure Duration Effects

Pinkerton et al. (1998); 

Harkema et al. (1997a); 

Harkema et al. (1997b); 

Catalano et al. (1995b); 

Catalano et al. (1995a); 

(Chang et al., 1995); 

Pinkerton et al. (1995); 

Stockstill et al. (1995); 

Harkema et al. (1994); 

Last et al. (1994); Plopper 

et al. (1994)

Rat, male and 

female, Fischer F344, 

6-8 weeks old

0.12

0.5

1.0

6 h/day, 5 days/week for 

20 months

Effects similar to (or a model of) early fibrotic 

human disease were greater in the periacinar 

region than in terminal bronchioles. Thickened 

alveolar septa observed at 0.12 ppm O3. Other 

effects (e.g., mucous cell metaplasia in the nose, 

mild fibrotic response in the parenchyma, and 

increased collagen in CAR of females) observed 

at 0.5 to 1.0 ppm. Some morphometric changes 

(epithelial thickening and bronchiolarization) 

occurred after 2 or 3 months of exposure to 

1.0 ppm.

Herbert et al. (1996) Mice, male and 

female, B6C3F1, 6-7 

weeks old,

0.12

0.50

1.0

6 h/day, 5 days/week for 

24 and 30 months

Similar to the response of rats in the same study 

(see rat above). Effects were seen in the nose and 

centriacinar region of the lung at 0.5 and 1.0 ppm.

Chang et al. (1991) Rat, male, F344, 

6 weeks old

Continuous: 0.12 

or 0.25

Episodic/urban: 

baseline 0.06; 

peak 0.25

Continuous: 12 h/day for 

6 weeks

Simulated urban pattern; 

slow rise to peak 9 

h/day, 5 days/week, 

13 weeks

Increased Type 1 and 2 epithelial volume 

assessed by TEM. Linear relationship observed 

between increases in Type 1 epithelial cell volume 

and concentration x time product. Degree of injury 

not related to pattern of exposure (continuous or 

episodic).

Chang et al. (1992) Rat, male, F344, 

6 weeks old

baseline 0.06; 

peak 0.25

Slow rise to peak 

9 h/day, 5 days/week, 

13 and 78 weeks

Recovery in filtered 

air for 6 or 17 weeks

Progressive epithelial hyperplasia, fibroblast 

proliferation, and interstitial matrix accumulation 

observed using TEM. Interstitial matrix thickening 

due to deposition of basement membrane and 

collagen fibers. Partial recovery of interstitial 

matrix during follow-up periods in air; but no 

resolution of basement membrane thickening.

Barry et al. (1985); (1983) Rat, male, 1 day old 

or 6 weeks old

0.12 (adults only)

0.25 

12 h/day for 6 weeks Lung and alveolar development not significantly 

affected. Increased Type 1 and 2 epithelial cells 

and AM in CAR alveoli, thickened Type 1 cells 

with smaller volume and less surface coverage as 

assessed by TEM (adults and juveniles). In adults, 

smaller but statistically significant similar changes 

at 0.12 ppm, suggesting linear concentrationresponse relationship. No statistically significant 

age-related effects observed.

Tyler et al. (1988) Monkey; male, 

Macaca fascicularis, 

7 mo old

0.25 8 h/day, 7 days/week, 

Daily for 18 mo or 

episodically every other 

month for 18 mo

Episodic group 

evaluated 1 mo 

postexposure

Increased collagen content, chest wall 

compliance, and inspiratory capacity in episodic 

group only. Respiratory bronchiolitis in both 

groups. Episodically exposed group incurred 

greater alterations in physiology and biochemistry 

and equivalent changes in morphometry even 

though exposed for half the time as the daily 

exposure group. 

Harkema et al. (1999) Rat, male, Fischer 

F344/N HSD, 10-14 

weeks old

0.25

0.5

8 h/day, 7 days/week for 

13 weeks

Mucous cell hyperplasia in nasal epithelium after 

exposure to 0.25 and 0.5 ppm O3; still evident 

after 13 weeks recovery from 0.5 ppm O3

exposure.

Van Bree et al. (2002) Rat, male, Wistar, 7 

weeks old, 

n = 5/group

0.4 23.5 h/day for 1, 3, 7, 

28,or 56 days

Acute inflammatory response in BALF reached a 

maximum at day 1 and resolved within 6 days 

during exposure. Centriacinar region inflammatory 

responses throughout O3 exposure with increased 

collagen and bronchiolization still present after a 

recovery period.

Katre et al. (2011) Mice; male, C57BL/6, 

6-8 weeks old

0.5 8 h/day, [5 days/week 

O3, and 2 days filtered 

air] for 5 or 10 cycles

Sustained elevation in TGF-β and PAI-1 in lung 

(5 or 10 cycles); elevated α-SMA and increased 

collagen deposition in airway walls (after 10 

cycles). Collagen increase shown to depend on 

TGF-β.
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Study Model O3 (ppm) Exposure Duration Effects

Schelegle et al. (2003); Monkey; Rhesus, 

30 days olda

0.5 8 h/day for 5 days, every 

5 days for a total of 11 

episodes

Goblet cell metaplasia, increased AHR, and 

increased markers of allergic asthma 

(e.g., eosinophilia) were observed, suggesting that 

episodic exposure to O3 alters postnatal 

morphogenesis and epithelial differentiation and 

enhances the allergic effects of house dust mite 

allergen in the lungs of infant primates.

Harkema et al. (1993, 

1987b)

Monkey; Macaca 

radiata, M, F

2-6 years old

0.15

0.3

8 h/day for 90 days Significant increase in epithelial thickness in 

respiratory bronchioles which was accompanied 

by increase in cuboidal cells; nasal lesions 

consisted of ciliated cell necrosis and secretory 

cell hyperplasia; no concentration response 

effects

Larson et al. (2004) Monkey; Macaca 

mulatta, 30 days olda

0.5 11 episodes of 5 days 

each, 8 h/day followed 

by 9 days of recovery

O3 or O3 + house dust mite antigen caused 

changes in density and number of airway 

epithelial nerves in small conducting airways. 

Suggests episodic O3 alters pattern of neural 

innervation in epithelial compartment of 

developing lungs.

Plopper et al. (2007) Monkey; Rhesus, 

30 days olda

0.5 5 months of episodic 

exposure; 5 days O3

followed by 9 days of 

filtered air, 8h/day. 

Non-significant increases airway resistance and 

airway responsiveness with O3 or inhaled allergen 

alone. Allergen + O3 produced additive changes in 

both measures. 

Fanucchi et al. (2006) Monkey; male 

Rhesus,30 days old

0.5 5 months of episodic 

exposure; 5 days O3

followed by 9 days of 

filtered air, 8h/day.

Cellular changes and significant structural 

changes in the distal respiratory tract in infant 

rhesus monkeys exposed to O3 postnatally.

Reiser et al. (1987) Monkey; male and 

female Cynomolgus 

6-7 mo old

0.61 8 h/day for 1 year Increased lung collagen content associated with 

elevated abnormal cross links that were 

irreversibly deposited.

a

sex not reported

1 Collectively, evidence from animal studies strongly suggests that chronic O3 exposure is 

2 capable of damaging the distal airways and proximal alveoli, resulting in lung tissue 

3 remodeling and leading to apparent irreversible changes. Potentially, persistent 

4 inflammation and interstitial remodeling play an important role in the progression and 

5 development of chronic lung disease. Further discussion of the modes of action that lead 

6 to O3-induced morphological changes can be found in Section 5.3.7. The findings 

7 reported in chronic animal studies offer insight into potential biological mechanisms for 

8 the suggested association between seasonal O3 exposure and reduced lung function 

9 development in children as observed in epidemiologic studies (see Section 7.2.3). 

10 Discussion of mechanisms involved in lifestage susceptibility and developmental effects 

11 can be found in Section 5.4.2.4.

7.2.4 Pulmonary Inflammation, Injury, and Oxidative Stress

12 The 2006 O3 AQCD stated that the extensive human clinical and animal toxicological 

13 evidence, together with the limited epidemiologic evidence available, suggests a causal 

14 role for O3 in inflammatory responses in the airways. Short-term exposure epidemiologic 
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1 studies discussed earlier in Section 6.2.3.2 show consistent associations of O3 exposure 

2 and increased airway inflammation and oxidative stress. Further discussion of the 

3 mechanisms underlying inflammation and oxidative stress responses can be found in 

4 Section 5.3.3. Though the majority of recent studies focus on short-term exposures, 

5 several epidemiologic and toxicology studies of long-term exposure add to observations 

6 of O3-induced inflammation and injury.

7 Inflammatory markers and peak expiratory pulmonary function were examined in 37 

8 allergic children with physician-diagnosed mild persistent asthma in a highly polluted 

9 urban area in Italy and then again 7 days after relocation to a rural location with 

10 significantly lower pollutant levels (Renzetti et al., 2009). The authors observed a 4-fold 

11 decrease in nasal eosinophils and a statistically significant decrease in fractional exhaled 

12 nitric oxide along with an improvement in lower airway function. Several pollutants were 

13 examined, including PM10, NO2, and O3, though pollutant-specific results were not 

14 presented. These results are consistent with studies showing that traffic-related exposures 

15 are associated with increased airway inflammation and reduced lung function in children 

16 with asthma and contribute to the notion that this negative influence may be rapidly 

17 reversible. Exhaled NO (eNO) has been shown to be a useful biomarker for airway 

18 inflammation in large population-based studies (Linn et al., 2009). Thus, while the time 

19 scale of 7 days between examinations for eNO needs to be evaluated for appropriateness, 

20 the results suggest that inflammatory responses are reduced when O3 levels are decreased.

21 Chest radiographs (CXR) of 249 children in Mexico City who were chronically exposed 

22 to O3 and PM2.5 were analyzed by Calderón-Garcidueñas et al. (2006). They reported an 

23 association between chronic exposures to O3 and other pollutants and a significant 

24 increase in abnormal CXR’s and lung CTs suggestive of a bronchiolar, peribronchiolar, 

25 and/or alveolar duct inflammatory process, in clinically healthy children with no risk 

26 factors for lung disease. These CXR and CT results should be viewed with caution 

27 because it is difficult to attribute effects to O3 exposure.

28 In a cross-sectional study, Wood et al. (2009) examined the association of outdoor air 

29 pollution with respiratory phenotype (PiZZ type) in alpha 1-antitrypsin deficiency 

30 (α-ATD) from the U.K. α-ATD registry. This deficiency leads to exacerbated responses 

31 to inflammatory stimuli. In total, 304 PiZZ subjects underwent full lung function testing 

32 and quantitative high-resolution computed tomography to identify the presence and 

33 severity of COPD – emphysema. Mean annual air pollution data for 2006 was matched to 

34 the location of patients’ houses and used in regression models to identify phenotypic 

35 associations with pollution controlling for covariates. Relative trends in O3 levels were 

36 assessed to validate use of a single year’s data to indicate long-term exposure and 

37 validation; data showed good correlations between modeled and measured data (Stedman 
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1 and Kent, 2008). Regression models showed that estimated higher exposure to O3

2 exposure was associated with worse gas transfer and more severe emphysema, albeit 

3 accounting for only a small proportion of the lung function variability. This suggests that 

4 a gene-specific group demonstrates a long-term O3 exposure effect.

5 The similarities of nonhuman primates to humans make them attractive models in which 

6 to study the effects of O3 on the respiratory tract. The nasal mucous membranes, which 

7 protect the more distal regions of the respiratory tract, are susceptible to injury from O3. 

8 Carey et al. (2007) conducted a study of O3 exposure in infant rhesus macaques, whose 

9 nasal airways closely resemble that of humans. Monkeys were exposed either acutely for 

10 5 days (8 h/day) to 0.5 ppm O3, or episodically for several biweekly cycles alternating 

11 5 days of 0.5 ppm O3 with 9 days of filtered air (0 ppm O3), designed to mimic human 

12 exposure (70 days total). All monkeys acutely exposed to O3 had moderate to marked 

13 necrotizing rhinitis, with focal regions of epitheliar exfoliation, numerous infiltrating 

14 neutrophils, and some eosinophils. The distribution, character, and severity of lesions in 

15 episodically exposed monkeys were similar to that of acutely exposed animals. Neither 

16 group exhibited the mucous cell metaplasia proximal to the lesions, observed in adult 

17 monkeys exposed continuously to 0.3 ppm O3 in another study (Harkema et al., 1987a). 

18 Adult monkeys also exhibit attenuation of inflammatory responses with continued daily 

19 exposure (Harkema et al., 1987a), but inflammation did not resolve over time in young 

20 episodically exposed monkeys (Carey et al., 2011). Inflammation in conducting airways 

21 has also been observed in rats chronically exposed to O3. Using an agar-based technique 

22 to fill the alveoli so that only the rat bronchi are lavaged, a 90-day exposure of rats to 

23 0.8 ppm O3 (8 h/day) elicited significantly elevated pro-inflammatory eicosanoids PGE2

24 and 12-HETE in the conducting airway compared to filtered air-exposed rats (Schmelzer 

25 et al., 2006).

26 Persistent inflammation and injury leading to interstitial remodeling may play an 

27 important role in the progression and development of chronic lung disease. Chronic 

28 airway inflammation is an important component of both asthma and COPD. The 

29 epidemiological evidence supporting an association between long-term exposure to O3

30 and inflammation or injury is limited. However, animal studies clearly demonstrate O3-

31 induced inflammation and injury, which may or may not attenuate with chronic exposure 

32 depending on the model. Further discussion of how O3 initiates inflammation can be 

33 found in Section 5.3.3. 
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7.2.5 Allergic Responses

1 The association of air pollutants with childhood respiratory allergies was examined in the 

2 U.S. using the 1999-2005 National Health Interview Survey of approximately 70,000 

3 children, and ambient air pollution data from the U.S. EPA, with monitors within 20 

4 miles of each child’s residential block (Parker et al., 2009). The authors examined the 

5 associations between the reporting of respiratory allergy or hay fever and medium-term 

6 exposure to O3 over several summer months, controlling for demographic and geographic 

7 factors. Increased respiratory allergy/hay fever was associated with increased O3 levels 

8 (adjusted OR per 10 ppb = 1.20; [95% CI: 1.15, 1.26]). These associations persisted after 

9 stratification by urban-rural status, inclusion of multiple pollutants (O3, SO2, NO2, PM), 

10 and definition of exposure by differing exposure radii; smaller samples within 5 miles of 

11 monitors were remarkably similar to the primary results. No associations between the 

12 other pollutants and the reporting of respiratory allergy/hay fever were apparent. 

13 Ramadour et al. (2000) reported no relationship between O3 levels and rhinitis symptoms 

14 and hay fever. Hwang et al. (2006) report the prevalence of allergic rhinitis (adjusted OR 

15 per 10 ppb = 1.05; [95% CI: 0.98, 1.12]) in a large cross-sectional study in Taiwan. In a 

16 large cross-sectional study in France, Penard-Morand et al. (2005) reported a positive 

17 relationship between lifetime allergic rhinitis and O3 exposure in a two-pollutant model 

18 with NO2. These studies related positive outcomes of allergic response and O3 exposure 

19 but with variable strength for the effect estimates. A toxicological study reported that 

20 five weeks of continuous exposure to 0.4 ppm O3 (but not 0.1 or 0.2 ppm O3) augmented 

21 sneezing and nasal secretions in a guinea pig model of nasal allergy (Iijima and 

22 Kobayashi, 2004). Nasal eosinophils, which participate in allergic disease and 

23 inflammation, and allergic antibody levels in serum were also elevated by exposure to 

24 concentrations as low as 0.2 ppm (Iijima and Kobayashi, 2004). 

25 Nasal eosinophils were observed to decrease by 4-fold in 37 atopic, mildly asthmatic 

26 children 7 days after relocation from a highly polluted urban area in Italy to a rural 

27 location with significantly lower pollutant levels (Renzetti et al., 2009). Inflammatory 

28 and allergic effects of O3 exposure (30 day mean) such as increased eosinophil levels 

29 were observed in children in an Austrian study (Frischer et al., 2001). Episodic exposure 

30 of infant rhesus monkeys to 0.5 ppm O3 for 5 months appears to significantly increase the 

31 number and proportion of eosinophils in the blood and airways (lavage) [protocol 

32 described above in Section 7.2.3.1 for Fanucchi et al. (2006)] (Maniar-Hew et al., 2011). 

33 These changes were not evident at 1 year of age (6 months after O3 exposure ceased). 

34 Increased eosinophils levels have also been observed after acute or prolonged exposures 

35 to O3 in adult bonnet and rhesus monkeys (Hyde et al., 1992; Eustis et al., 1981). 
Draft – Do Not Cite or Quote 7-33 June 2012

1 Total IgE levels were related to air pollution levels in 369 adult asthmatics in five French 

2 centers using generalized estimated equations (GEE) as part of the EGEA study described 

3 earlier (Rage et al., 2009a). Geostatistical models were performed on 4×4 km grids to 

4 assess individual outdoor air pollution exposure that was assigned to subject’s home 

5 address. Ozone concentrations were positively related to total IgE levels and an increase 

6 of 5 ppb of O3 resulted in an increase of 20.4% (95% CI: 3.0, 40.7) in total IgE levels. 

7 Nearly 75% of the subjects were atopic. In two-pollutant models including O3 and NO2, the 

8 O3 effect estimate was decreased by 25% while the NO2 effect estimate was decreased by 

9 57%. Associations were not sensitive to adjustment for covariates or the season of IgE 

10 measurements. These cross-sectional results suggest that exposure to O3 may increase 

11 total IgE in adult asthmatics. 

12 Although very few toxicological studies of long-term exposure examining allergy are 

13 available, short-term exposure studies in rodents and nonhuman primates demonstrate 

14 allergic skewing of immune responses and enhanced IgE production. Due to the 

15 persistent nature of these responses, the short-term toxicological evidence lends 

16 biological plausibility to the limited epidemiologic findings of an association between 

17 long-term O3 exposure and allergic outcomes.

7.2.6 Host Defense

18 Short-term exposures to O3 have been shown to cause decreases in host defenses against 

19 infectious lung disease in animal models. Acute O3-induced suppression of alveolar 

20 phagocytosis and immune functions observed in animals appears to be transient and 

21 attenuated with continuous or repeated exposures, although chronic exposure (weeks, 

22 months) has been shown to slow alveolar clearance. In an important study investigating 

23 the effects of longer term O3 exposure on alveolobronchiolar clearance, rats were exposed 

24 to an urban pattern of O3 (continuous 0.06 ppm, 7 days/week with a slow rise to a peak of 

25 0.25 ppm and subsequent decrease to 0.06 ppm over a 9 h period for 5 days/week) for 

26 6 weeks and were exposed 3 days later to chrysotile asbestos, which can cause pulmonary 

27 fibrosis and neoplasia (Pinkerton et al., 1989). After 30 days, the lungs of the O3-exposed 

28 animals had twice the number and mass of asbestos fibers as the air-exposed rats. 

29 However, chronic exposures of 0.1 ppm do not cause greater effects on infectivity than 

30 short exposures, due to defense parameters becoming reestablished with prolonged 

31 exposures. No detrimental effects were seen with a 120-day exposure to 0.5 ppm O3 on 

32 acute lung injury from influenza virus administered immediately before O3 exposure 

33 started. However, O3 was shown to increase the severity of postinfluenzal alveolitis and 

34 lung parenchymal changes (Jakab and Bassett, 1990). A recent study by Maniar-Hew et 

35 al. (2011) demonstrated that the immune system of infant rhesus monkeys episodically 
Draft – Do Not Cite or Quote 7-34 June 2012

exposed to 0.5 ppm O3 for 5 months1 1 appeared to be altered in ways that could diminish 

2 host defenses. Reduced numbers of circulating leukocytes were observed, particularly 

3 polymorphonuclear leukocytes (PMNs) and lymphocytes, which were decreased in the 

4 blood and airways (bronchoalveolar lavage). These changes did not persist at 1 year of 

5 age (6 months postexposure); rather, increased numbers of monocytes were observed at 

6 that time point. Challenge with LPS, a bacterial ligand that activates monocytes and other 

7 innate immune cells, elicited lower responses in O3-exposed animals even though the 

8 relevant reactive cell population was increased. This was observed in both an in vivo 

9 inhalation challenge and an ex vivo challenge of peripheral blood mononuclear cells. 

10 Thus a decreased ability to respond to pathogenic signals was observed six months after 

11 O3 exposure ceased, in both the lungs and periphery.

7.2.7 Respiratory Mortality

12 A limited number of epidemiologic studies have assessed the relationship between long13 term exposure to O3 and mortality. The 2006 O3 AQCD concluded that an insufficient 

14 amount of evidence existed “to suggest a causal relationship between chronic O3

15 exposure and increased risk for mortality in humans” (U.S. EPA, 2006b). Though total 

16 and cardio-pulmonary mortality were considered in these studies, respiratory mortality 

17 was not specifically considered. In the most recent follow-up analysis of the ACS cohort 

18 (Jerrett et al., 2009), cardiopulmonary deaths were subdivided into respiratory and 

19 cardiovascular, separately, as opposed to combined in the Pope et al. (2002) work. A 

20 10-ppb increment in exposure to O3 elevated the risk of death from respiratory causes and 

21 this effect was robust to the inclusion of PM2.5. The association between increased O3

22 concentrations and increased risk of death from respiratory causes was insensitive to the 

23 use of a random-effects survival model allowing for spatial clustering within the 

24 metropolitan area and state of residence, and to adjustment for several ecologic variables 

25 considered individually. Additionally, a recent study (Zanobetti and Schwartz, 2011)

26 observed an association between long-term exposure to O3 and elevated risk of mortality 

27 among Medicare enrollees that had previously experienced an emergency hospital 

28 admission due to COPD.

7.2.8 Summary and Causal Determination

29 The epidemiologic studies reviewed in the 2006 O3 AQCD detected no associations 

30 between long-term (annual) O3 exposures and asthma-related symptoms, asthma 

 1 Exposure protocol is described above in Section 7.2.3.2 for Fanucchi et al. (2006).
Draft – Do Not Cite or Quote 7-35 June 2012

1 prevalence, or allergy to common aeroallergens among children after controlling for 

2 covariates. Little evidence was available to relate long-term exposure to ambient O3

3 concentrations with deficits in the growth rate of lung function in children. Additionally, 

4 limited evidence was available evaluating the relationship between long-term O3

5 concentrations and pulmonary inflammation and other endpoints. From toxicological 

6 studies, it appeared that O3-induced inflammation tapered off during long-term 

7 exposures, but that hyperplastic and fibrotic changes remained elevated and in some 

8 cases even worsened after a postexposure period in clean air. Episodic exposures were 

9 also known to cause more severe pulmonary morphologic changes than continuous 

10 exposure (U.S. EPA, 2006b).

11 The recent epidemiologic evidence base consists of studies using a variety of designs and 

12 analysis methods evaluating the relationship between long-term exposure to ambient O3

13 concentrations and measures of respiratory health effects and mortality conducted by 

14 different research groups in different locations. See Table 7-2 for O3 concentrations 

15 associated with selected studies. Table 7-2 is organized by longitudinal and cross16 sectional studies both presented alphabetically. The positive results from various designs 

17 and locations support a relationship between long-term exposure to ambient O3

18 concentrations and respiratory health effects and mortality. 

19 Earlier studies reported associations of new-onset asthma and O3 in an adult cohort in 

20 California (McDonnell et al., 1999a; Greer et al., 1993) but only in males. In the CHS 

21 cohort of children in 12 Southern California communities, long-term exposure to O3

22 concentrations was not associated with increased risk of developing asthma (McConnell 

23 et al., 2010); however, greater outdoor exercise was associated with development of 

24 asthma in children living in communities with higher ambient O3 concentrations 

25 (McConnell et al., 2002). Recent CHS studies examined interactions among genetic 

26 variants, long-term O3 exposure, and new onset asthma in children. These prospective 

27 cohort studies are methodologically rigorous epidemiology studies, and evidence 

28 indicates gene-O3 interactions. These studies have provided data supporting decreased 

29 risk of certain different genetic variants on new onset asthma (e.g., HMOX-1, ARG) that 

30 is limited to children either in low (Islam et al., 2008) or high (Salam et al., 2009) O3

31 communities. Gene-environment interaction also was demonstrated with findings that 

32 greater outdoor exercise increased risk of asthma in GSTP1 Ile/Ile children living in high 

33 O3 communities (Islam et al., 2009). Biological plausibility for these these gene-O3

34 environment interactions is provided by evidence that these enzymes have antioxidant 

35 and/or anti-inflammatory activity and participate in well recognized modes of action in 

36 asthma pathogenesis. As O3 is a source of oxidants in the airways, oxidative stress serves 

37 as the link among O3 exposure, enzyme activity, and asthma. 
Draft – Do Not Cite or Quote 7-36 June 2012

Table 7-2 Summary of selected key new studies examining annual ozone 

exposure and respiratory health effects

Study; 

Health Effect; 

Location Annual Mean O3 Concentration (ppb)

O3 Range 

(ppb) 

Percentiles

Longitudinal

Islam et al. (2008); 

New-onset asthma; 

CHS

55.2 high vs. 38.4 low communities 

10:00 a.m. to 6:00 p.m. average

See left

Islam et al. (2009); 

New-onset asthma; 

CHS

55.2 high vs. 38.4 low communities 

10:00 a.m. to 6:00 p.m.

See left

Lin et al. (2008b); 

First asthma hospital admission; 

New York State - 10 regions

Range of mean O3 concentrations over the 

10 New York Regions 37.51 to 47.78

8-h max 10:00 a.m. to 6:00 p.m.

See left

Salam et al. (2009); 

Childhood onset asthma; 

CHS

O3 greater than or less than 50 ppb See left

Cross-sectional

Akinbami et al. (2010); 

Current asthma

United States

12 month median 39.8

8hr max

IQR 

35.9 to 43.7

Hwang et al. (2005); 

Prevalence of asthma; 

Taiwan

Mean 23.14 Range 

18.65 to 31.17

Jacquemin et al. (In Press); 

Asthma control in adults; 

Five French cities

Median 46.9 ppb;

8-h average

25th-75th

41-52

.Lee et al. (2009b); 

Bronchitic symptoms in asthmatic children; 

CHS

Above and below 50 ppb See left

Meng et al. (2010); 

Asthma ED visits or hospitalizations; 

San Joaquin Valley, CA

Median 30.3 ppb

Yearly based on hourly

25-75% range 

27.1 to 34.0

Moore et al. (2008); 

Asthma hospital admissions; 

South Coast Basin

Median 87.8 ppb

Quarterly 1hr daily max

Range 

28.6 to 199.9

Rage et al. (2009a); 

Asthma severity; 

Five French cities

Mean 30 ppb

8-h average

25th-75th 

21-36

Wenten et al. (2009); 

Respiratory school absence, 

U.S.

Median 46.9 ppb; 

10a.m. – 6 p.m. average

Min-Max 

27.6-65.3

1 Studies using a cross-sectional design provide support for a relationship between long2 term O3 exposure and health effects in asthmatics. A long-term O3 exposure study relates 

3 bronchitic symptoms to TNF-308 genotype asthmatic children with ambient O3 exposure 

4 in the CHS (Lee et al., 2009b). A study in five French cities reports effects on asthma 

5 severity related to long-term O3 exposure (Rage et al., 2009b). A follow-up study of this 
Draft – Do Not Cite or Quote 7-37 June 2012

1 cohort (Jacquemin et al., In Press) supports an effect of cumulative long-term O3

2 exposure on asthma control in adulthood in subjects with pre-existing asthma. Akinbami 

3 et al. (2010) and Hwang et al. (2005) provide further evidence relating O3 exposures and 

4 the risk of asthma. For the respiratory health of a cohort based on the general U.S. 

5 population, risk of respiratory-related school absences was elevated for children with the 

6 CAT and MPO variant genes related to communities with high ambient O3 levels 

7 (Wenten et al., 2009). 

8 Long-term O3 exposure was related to first childhood asthma hospital admissions in a 

9 positive concentration-response relationship in a New York State birth cohort (Lin et al., 

10 2008b). A separate hospitalization cross-sectional study in San Joaquin Valley, California 

11 reports similar findings (Meng et al., 2010). Another study relates asthma hospital 

12 admissions to quarterly average O3 in the South Coast Air Basin of California (Moore et 

13 al., 2008).

14 Information from toxicological studies indicates that long term exposure to O3 during 

15 gestation or development can result in irreversible morphological changes in the lung, 

16 which in turn can influence the function of the respiratory tract. Studies by Plopper and 

17 colleagues using an allergic asthma model have demonstrated changes in pulmonary 

18 function and airway morphology in adult and infant nonhuman primates repeatedly 

19 exposed to environmentally relevant concentrations of O3 (Fanucchi et al., 2006; Joad et 

20 al., 2006; Schelegle et al., 2003; Harkema et al., 1987b). This nonhuman primate 

21 evidence of an O3-induced change in airway responsiveness supports the biologic 

22 plausibility of long term exposure to O3 contributing to effects of asthma in children. 

23 Results from epidemiologic studies examining long-term O3 exposure and pulmonary 

24 function effects are inconclusive with some new studies relating effects at higher 

25 exposure levels. The definitive 8-year follow-up analysis of the first cohort of the CHS, 

26 which is discussed in Section 7.2 (Gauderman et al., 2004), provided little evidence that 

27 long-term exposure to ambient O3 was associated with significant deficits in the growth 

28 rate of lung function in children. Other cross-sectional studies provide mixed results. 

29 Several studies (see Table 7-3) provide results adjusted for potential confounders, 

30 presenting results for both O3 and PM (single and multipollutant models) as well as other 

31 pollutants where PM effects were not provided. As shown in the table, O3 associations 

32 are generally robust to adjustment for potential confounding by PM.
Draft – Do Not Cite or Quote 7-38 June 2012

Table 7-3 Studies providing evidence concerning potential confounding by 

PM for available endpoints.

Study

Endpoint Exposure

Single 

Pollutant O3

Single Pollutant 

PM O3 with PM PM with O3

Asthma Related Health Effect Endpoint

Akinbami et al. (2010)

Asthma prevalence in 

children

IQR 

35.9-43.7 ppb

1.56

(1.15, 2.10)

PM2.5

1.43

(0.98, 2.10)

Adjusted for 

SO2,PM2.5,PM10

1.86 (1.02-3.40)

Adjusted for 

PM2.5, PM10

1.36 (0.91-2.02)

PM2.5

1.24 

(0.70-2.21)

PM2.5

1.26

0.80-1.98)

Hwang et al. (2005)

Asthma risk in children

10 ppb O3 1.138

(1.001, 1.293

0.934

(0.909, 0.960)

PM10

1.253 

(1.089, 1.442)

0.925

(0.899, 0.952)

Jacquemin et al. (In 

Press)

Asthma control in adults

IQR 25-38 ppb O3

summer

1.69

(1.22, 2.34)

1.33

(1.06, 1.67)

PM10

1.50 

(1.07, 2.11)

1.28

(1.06, 1.55)

Lee et al. (2009b)

Bronchitic symptoms 

asthmatics

High O3 >50 ppb 1.42

(0.75, 2.70)

NA No substantial 

differences

PM10, PM2.5

NA

Lin et al. (2008b)

Asthma admissions in 

children

IQR 2.5% 1.16

(1.15, 1.17)

NA Air Quality Index

1.24 

(1.23, 1.25)

NA

Meng et al. (2007)

Asthma control 

1 ppm 1.70

(0.91, 3.18)

PM10

2.06

(1.17, 3.61)

women

Did not differ NA

Meng et al. (2010)

Asthma ED visits, 

Hospitalization

10 ppb 1.49

(1.05, 2.11)

PM10

1.29

(0.99, 1.69)

Did not differ NA

Rage et al. (2009b)

Asthma severity in adults

IQR 

28.5-33.9 ppb

2.53

(1.69, 3.79)

NA No PM data

Three pollutant 

(O3, NO2, SO2)

2.74 

(1.68, 4.48)

NA

Other Respiratory Health Effect Endpoints

Karr et al. (2007)

Bronchiolitis

Hospitalization

10 ppb 0.92

(0.88, 0.96)

1.09

(1.04, 1.14)

PM2.5

1.02 

(0.94, 1.10)

1.09

(1.03, 1.15)

Parker et al. (2009)

Respiratory allergy

10 ppb 1.24

(1.15, 1.34)

1.23

(1.04, 1.46)

Multi-pollutant

1.18 

(1.09, 1.27)

1.29 (

1.07, 1.56)

Rojas-Martinez et al. 

(2007)

FEV1 (mL) Deficit

Girls

11.3 ppb IQR -24

(-30, -19)

PM10 IQR

36.4 ug/m3

-29(-36, -21)

-17 

(-23, -12)

-24 

(-31,-16)

The highest quartile is shown for all results

NA = not available
Draft – Do Not Cite or Quote 7-39 June 2012

1 There is limited evidence for an association between long-term exposure to ambient O3

2 concentrations and respiratory mortality (Jerrett et al., 2009) and this effect was robust to 

3 the inclusion of PM2.5. The association between increased O3 concentrations and 

4 increased risk of death from respiratory causes was insensitive to a number of different 

5 model specifications. Additionally, there is evidence that long-term exposure to O3 is 

6 associated with mortality among individuals that had previously experienced an 

7 emergency hospital admission due to COPD (Zanobetti and Schwartz, 2011).

8 Taken together, the recent epidemiologic studies of respiratory health effects (including 

9 respiratory symptoms, new-onset asthma and respiratory mortality) combined with 

10 toxicological studies in rodents and nonhuman primates, provide biologically plausible 

11 evidence that there is likely to be a causal relationship between long-term exposure 

12 to O3 and respiratory effects. The strongest epidemiologic evidence for a relationship 

13 between long-term O3 exposure and respiratory effects is provided by studies that 

14 demonstrate interactions between exercise or different genetic variants and long-term 

15 measures of O3 exposure on new-onset asthma in children; and increased respiratory 

16 symptom effects in asthmatics. Additional studies of respiratory health effects and a 

17 study of respiratory mortality provide a collective body of evidence supporting these 

18 relationships. Studies considering other pollutants provide data suggesting that the effects 

19 related to O3 are independent from potential effects of the other pollutants. Some studies 

20 provide evidence for a positive concentration-response relationship. Short-term studies 

21 provide supportive evidence with increases in respiratory symptoms and asthma 

22 medication use, hospital admissions and ED visits for all respiratory outcomes and 

23 asthma, and decrements in lung function in children. The recent epidemiologic and 

24 toxicological data base provides a compelling case to support the hypothesis that a 

25 relationship exists between long-term exposure to ambient O3 and measures of 

26 respiratory health effects. 

7.3 Cardiovascular Effects

7.3.1 Cardiovascular Disease

7.3.1.1 Cardiovascular Epidemiology

27 Long-term exposure to O3 and its effects on cardiovascular morbidity were not 

28 considered in the 2006 O3 AQCD. However, recent studies have assessed the chronic 

29 effects of O3 concentration on cardiovascular morbidity (Chuang et al., 2011; Forbes et 
Draft – Do Not Cite or Quote 7-40 June 2012

1 al., 2009a; Chen et al., 2007a). The association between O3 concentration and markers of 

2 lipid peroxidation and antioxidant capacity was examined among 120 nonsmoking 

3 healthy college students, aged 18-22 years, from the University of California, Berkeley 

4 (February—June 2002) (Chen et al., 2007a). By design, students were chosen from 

5 geographic areas so they had experienced different concentrations of O3 over their 

6 lifetimes and during recent summer vacation in either greater Los Angeles (LA) or the 

7 San Francisco Bay Area (SF). A marker of lipid peroxidation, 8-isoprostane (8-iso-PGF) 

8 in plasma, was assessed. This marker is formed continuously under normal physiological 

9 conditions but has been found at elevated concentrations in response to environmental 

10 exposures. A marker of overall antioxidant capacity, ferric reducing ability of plasma 

11 (FRAP), was also measured. The lifetime average O3 concentration estimates (from 

12 estimated monthly averages) did not show much overlap between the two geographic 

13 areas [median (range): LA, 42.9 ppb (28.5-65.3); SF, 26.9 ppb (17.6-33.5)]. Estimated 

14 lifetime average O3 concentration was related to 8-iso-PGF [β = 0.025 (pg/mL)/8-h ppb 

15 O3, p = 0.0007]. For the 17-ppb lifetime O3 concentration difference between LA and SF 

16 participants, there was a 17.41-pg/mL (95% CI: 15.43, 19.39) increase in 8-iso-PGF. No 

17 evidence of association was observed between lifetime O3 concentration and FRAP 

18 [β = -2.21 (pg/mL)/8-h ppb O3, p = 0.45]. The authors note that O3 was highly correlated 

19 with PM10-2.5 and NO2 in this study population; however, their inclusion in the O3 models 

20 did not substantially modify the magnitude of the associations with O3. Because the 

21 average lifetime concentration results were supported by shorter-term exposure period 

22 results from analyses considering O3 concentrations up to 30 days prior to sampling, the 

23 authors conclude that persistent exposure to O3 can lead to sustained oxidative stress and 

24 increased lipid peroxidation. However, because there was not much overlap in average 

25 lifetime O3 concentration estimates between LA and SF, it is possible that the risk 

26 estimates involving the lifetime O3 exposures could be confounded by unmeasured 

27 factors related to other differences between the two cities. 

28 Forbes et al. (2009a) used the annual average exposures to assess the relationship 

29 between chronic ambient air pollution and levels of fibrinogen and C-reactive protein 

30 (CRP) in a cross-sectional study conducted in England. Data were collected from the 

31 Health Survey of England for 1994, 1998, and 2003. The sampling strategy was designed 

32 to obtain a representative sample of the English population; however, due to small group 

33 sizes, only data from white ethnic groups were analyzed. For analyses, the annual 

34 concentrations of O3 were averaged for the year of data collection and the previous year 

35 with the exception of 1994 (because pollutant data were not available for 1993). Median 

36 O3 concentrations were 26.7 ppb, 25.4 ppb, and 28 ppb for 1994, 1998, and 2003, 

37 respectively. Year specific adjusted effect estimates were created and combined in a 

38 meta-analysis. No evidence of association was observed for O3 and levels of fibrinogen 

39 or CRP (e.g., the combined estimates for the percent change in fibrinogen and CRP for a 
Draft – Do Not Cite or Quote 7-41 June 2012

1 10 ppb increase in O3 were -0.28 [95% CI: -2.43, 1.92] and -3.05 [95% CI: -16.10, 

2 12.02], respectively). 

3 A study was performed in Taiwan to examine the association between long-term O3

4 concentrations and blood pressure and blood markers using the Social Environment and 

5 Biomarkers of Aging Study (SEBAS) (Chuang et al., 2011). Individuals included in the 

6 study were 54 years of age and older. The mean annual O3 concentration during the study 

7 period was 22.95 ppb (SD 6.76 ppb). Positive associations were observed between O3

8 concentrations and both systolic and diastolic blood pressure [changes in systolic and 

9 diastolic blood pressure were 21.51mmHg (95% CI: 16.90, 26.13) and 20.56 mmHg 

10 (95% CI: 18.14, 22.97) per 8.95 ppb increase in O3, respectively). Increased O3

11 concentrations were also associated with increased levels of total cholesterol, fasting 

12 glucose, hemoglobin A1c, and neutrophils. No associations were observed between O3

13 concentrations and triglyceride and IL-6 levels. The observed associations were reduced 

14 when other pollutants were added to the models. Further research will be important for 

15 understanding the effects, if any, of chronic O3 exposure on cardiovascular morbidity 

16 risk.

7.3.1.2 Cardiovascular Toxicology

17 Three new studies have investigated the cardiovascular effects of long-term exposure to 

18 O3 in animal models (See Table 7-3 for study details). In addition to the short-term 

19 exposure effects described in Section 6.3.3, a recent study found that O3 exposure in 

20 genetically hyperlipidemic mice enhanced aortic atherosclerotic lesion area compared to 

21 air exposed controls (Chuang et al., 2009). Chuang et al. (2009) not only provided 

22 evidence for increased atherogenesis in susceptible mice, but also reported an elevated 

23 vascular inflammatory and redox state in wild-type mice and infant primates 

24 (Section 6.3.3). This study is compelling in that it identifies biochemical and cellular 

25 events responsible for transducing the airway epithelial reactions of O3 into 

26 proinflammatory responses that are apparent in the extrapulmonary vasculature (Cole and 

27 Freeman, 2009). 

28 Another recent study provides further evidence for increased vascular inflammation and 

29 oxidation and long term effects in the extrapulmonary space. Rats episodically exposed to 

30 O3 for 16 weeks presented marked increases in gene expression of biomarkers of 

31 oxidative stress, thrombosis, vasoconstriction, and proteolysis (Kodavanti et al., 2011). 

32 Ozone exposure upregulated aortic mRNA expression of heme oxygenase-1 (HO-1), 

33 tissue plasminogen activator (tPA), plasminogen activator inhibitor-1 (PAI-1), von 

34 Willebrand factor (vWf), thrombomodulin, endothelial nitric oxide synthase (eNOS), 
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1 endothelin-1 (ET-1), matrix metalloprotease-2 (MMP-2), matrix metalloprotease-3 

2 (MMP-3), and tissue inhibitor of matrix metalloprotease-2 (TIMP-2). In addition, O3

3 exposure depleted some cardiac mitochondrial phospholipid fatty acids (C16:0 and 

4 C18:1), which may be the result of oxidative modifications. The authors speculate that 

5 oxidatively modified lipids and proteins produced in the lung and heart promote vascular 

6 pathology through activation of lectin-like oxidized-low density lipoprotein receptor-1 

7 (LOX-1). Activated LOX-1 induces expression of a number of the biomarkers induced by 

8 O3 exposure and is considered pro-atherogenic. Both LOX-1 mRNA and protein were 

9 increased in mouse aorta after O3 exposure. This study provides a possible pathway and 

10 further support to the observed O3 induced atherosclerosis.

11 Vascular occlusion resulting from atherosclerosis can block blood flow through vessels 

12 causing ischemia. The restoration of blood flow or reperfusion can cause injury to the 

13 tissue from subsequent inflammation and oxidative damage. Ozone exposure enhanced 

14 the sensitivity to myocardial ischemia-reperfusion (I/R) injury in rats while increasing 

15 oxidative stress levels and pro-inflammatory mediators and decreasing production of 

16 anti-inflammatory proteins (Perepu et al., 2010). Both long- and short-term O3 exposure 

17 decreased the left ventricular developed pressure, rate of change of pressure 

18 development, and rate of change of pressure decay and increased left ventricular end 

19 diastolic pressure in isolated perfused hearts (Section 6.3.3 for short-term exposure 

20 discussion). In this ex vivo heart model, O3 induced oxidative stress by decreasing SOD 

21 enzyme activity and increasing malondialdehyde levels. Ozone also elicited a 

22 proinflammatory state evident by an increase in TNF-α and a decrease in the 

23 anti-inflammatory cytokine IL-10. The authors conclude that O3 exposure will result in a 

24 greater I/R injury.

25 Overall, the few animal studies that have been conducted suggest that long-term O3

26 exposure may result in cardiovascular effects. These studies demonstrate O3-induced 

27 atherosclerosis and injury. In addition, evidence is presented for a potential mechanism 

28 for the development of vascular pathology that involves increased oxidative stress and 

29 proinflammatory mediators, activation of LOX-1 by O3 oxidized lipids and proteins, and 

30 upregulation of genes responsible for proteolysis, thrombosis, and vasoconstriction. 

31 Further discussion of the mechanisms that may lead to cardiovascular effects from O3

32 exposure can be found in Section 5.3.8. 
Draft – Do Not Cite or Quote 7-43 June 2012

Table 7-3 Characterization of Study Details for Section 7.3.1.2.

Study Model O3 (ppm) Exposure Duration Effects

Chuang et al. (2009) Mice; ApoE-/-; M; 

6 weeks

0.5 8 wks, 5 days/week, 

8 h/day

Enhanced aortic atherosclerotic lesion 

area compared to air controls.

Kodavanti et al. 

(2011)

Rat; Wistar; M; 

10-12 weeks

0.4 16 wks, 1 day/week, 

5 h/day

Increased vascular inflammation and 

oxidative stress, possibly through 

activation of LOX-1 signaling.

Perepu et al. (2010) Rat; Sprague-Dawley; 

Weight: 50-75 g

0.8 56 days, 8 h/day Enhanced the sensitivity to myocardial I/R 

injury while increasing oxidative stress 

and pro-inflammatory mediators and 

decreasing production of 

anti-inflammatory proteins.

No previous studies investigated cardiovascular effects from long-term exposure to O3.

For details, see Section 7.3.1.2

7.3.2 Cardiovascular Mortality

1 A limited number of epidemiologic studies have assessed the relationship between long2 term exposure to O3 and mortality. The 2006 O3 AQCD concluded that an insufficient 

3 amount of evidence existed “to suggest a causal relationship between chronic O3

4 exposure and increased risk for mortality in humans” (U.S. EPA, 2006b). Though total 

5 and cardio-pulmonary mortality were considered in these studies, cardiovascular 

6 mortality was not specifically considered. In the most recent follow-up analysis of the 

7 ACS cohort (Jerrett et al., 2009), cardiopulmonary deaths were subdivided into 

8 respiratory and cardiovascular, separately, as opposed to combined in the Pope et al. 

9 (2002) work. A 10-ppb increment in exposure to O3 elevated the risk of death from the 

10 cardiopulmonary, cardiovascular, and ischemic heart disease. Inclusion of PM2.5 as a 

11 copollutant attenuated the association with exposure to O3 for all of the cardiovascular 

12 endpoints to become null. Additionally, a recent study (Zanobetti and Schwartz, 2011)

13 observed an association between long-term exposure to O3 and elevated risk of mortality 

14 among Medicare enrollees that had previously experienced an emergency hospital 

15 admission due to congestive heart failure (CHF) or myocardial infarction (MI).

7.3.3 Summary and Causal Determination

16 Previous AQCDs did not address the cardiovascular effects of long-term O3 exposure due 

17 to limited data availability. The evidence remains limited; however the emerging data is 

18 supportive of a role for O3 in chronic cardiovascular diseases. Few epidemiologic studies 

19 have investigated cardiovascular morbidity after long-term O3 exposure, and the majority 

20 only assessed cardiovascular disease related biomarkers. The studies used annual or 
Draft – Do Not Cite or Quote 7-44 June 2012

1 multi-year averages of air monitoring data for exposure assessment. As described in 

2 Section 4.6, this exposure assignment method is typical of long-term epidemiologic 

3 studies, and analyses suggest that annual average concentrations are representative of 

4 exposure metrics accounting for residential mobility. A study on O3 and cardiovascular 

5 mortality reported no association after adjustment for PM2.5 levels. Further epidemiologic 

6 studies on cardiovascular morbidity and mortality after long-term exposure have not been 

7 published. 

8 Toxicological evidence on long-term O3 exposure is also limited but three strong 

9 toxicological studies have been published since the previous AQCD. These studies 

10 provide evidence for O3 enhanced atherosclerosis and I/R injury, corresponding with 

11 development of a systemic oxidative, proinflammatory environment. Further discussion 

12 of the mechanisms that may lead to cardiovascular effects can be found in Section 5.3.8. 

13 Although questions exist for how O3 inhalation causes systemic effects, a recent study 

14 proposes a mechanism for development of vascular pathology that involves activation of 

15 LOX-1 by O3 oxidized lipids and proteins. This activation may also be responsible for O3

16 induced changes in genes involved in proteolysis, thrombosis, and vasoconstriction. 

17 Taking into consideration the findings of toxicological studies, and the emerging 

18 evidence from epidemiologic studies, the generally limited body of evidence is

19 suggestive of a causal relationship between long-term exposures to O3 and 

20 cardiovascular effects. 

7.4 Reproductive and Developmental Effects 

21 Although the body of literature characterizing the health effects associated with exposure 

22 to O3 is large and continues to grow, the research focusing on adverse birth outcomes is 

23 relatively small. Among these studies, various measures of birth weight and fetal growth, 

24 such as low birth weight (LBW), small for gestational age (SGA), and intrauterine 

25 growth restriction (IUGR), and preterm birth (<37-week gestation; [PTB]) have received 

26 more attention in air pollution research, while congenital malformations are less studied. 

27 There are also recent studies on reproductive and developmental effects and infant 

28 mortality.

29 A major issue in studying environmental exposures and reproductive and developmental 

30 effects (including infant mortality) is selecting the relevant exposure period, since the 

31 biological mechanisms leading to these outcomes and the critical periods of exposure are 

32 poorly understood. To account for this, many epidemiologic studies evaluate multiple 

33 exposure periods, including long-term (months to years) exposure periods, such as entire 

34 pregnancy, individual trimesters or months of pregnancy, and short-term (days to weeks) 
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1 exposure periods such as the days and weeks immediately preceding birth. Due to the 

2 length of gestation in rodents (18-24 days, on average), animal toxicological studies 

3 investigating the effects of O3 generally utilize short-term exposure periods. Thus, an 

4 epidemiologic study that uses the entire pregnancy as the exposure period is considered 

5 to have a long-term exposure period (about 40 weeks, on average), while a toxicological 

6 study conducted with rats that also uses the entire pregnancy as the exposure period is 

7 considered to have a short-term exposure period (about 18-24 days, on average). In order 

8 to characterize the weight of evidence for the effects of O3 on reproductive and 

9 developmental effects in a consistent, cohesive and integrated manner, results from both 

10 short-term and long-term exposure periods are included in this section and are identified 

11 accordingly in the text and tables throughout this section. 

12 Due to the poorly understood biological mechanisms and uncertainty regarding relevant 

13 exposure studies, all of the studies of reproductive and developmental outcomes, 

14 including infant mortality, are evaluated in this section. Infant development processes, 

15 much like fetal development processes, may be particularly sensitive to O3-induced 

16 health effects. Exposures proximate to the death may be most relevant if exposure causes 

17 an acute effect. However, exposure occurring in early life might affect critical growth and 

18 development, with results observable later in the first year of life, or cumulative exposure 

19 during the first year of life may be the most important determinant. In dealing with the 

20 uncertainties surrounding these issues, studies have considered several exposure metrics 

21 based on different periods of exposure, including both short- and long-term exposure 

22 periods. In the toxicological literature,a challenge in interpreting data from studies that 

23 use very young murine pups, is that pups can have differential exposure to O3 doses, 

24 versus their respective dams, because of the physiology and behavior associated with the 

25 early postnatal period. Namely, young pups tend to nuzzle close to their mothers and are 

26 often housed in cages with litter used in nest formation. Both the dam's fur and the 

27 bedding can absorb and react with O3, decreasing the dose that a young animal might 

28 receive. The reproductive and developmental studies are characterized in this chapter, as 

29 they contribute to the weight of evidence for an effect of O3 on reproductive and 

30 developmental effects.

31 Infants and fetal development processes may be particularly at-risk for O3-induced health 

32 effects, and although the physical mechanisms are not fully understood, several 

33 hypotheses have been proposed; these include: oxidative stress, systemic inflammation, 

34 vascular dysfunction and impaired immune function (Section 5.3). Study of these 

35 outcomes can be difficult given the need for detailed exposure data and potential 

36 residential movement of mothers during pregnancy. Air pollution epidemiologic studies 

37 reviewed in the 2006 O3 AQCD examined impacts on birth-related endpoints, including 

38 intrauterine, perinatal, postneonatal, and infant deaths; premature births; intrauterine 
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1 growth retardation; very low birth weight (weight <1,500 grams) and low birth weight 

2 (weight <2,500 grams); and birth defects. However, in the limited number of studies that 

3 investigated O3, no associations were found between O3 and birth outcomes, with the 

4 possible exception of birth defects. 

5 Several recent articles have reviewed methodological issues relating to the study of 

6 outdoor air pollution and adverse birth outcomes (Chen et al., 2010a; Woodruff et al., 

7 2009; Ritz and Wilhelm, 2008; Slama et al., 2008). Some of the key challenges to 

8 interpretation of these study results include the difficulty in assessing exposure as most 

9 studies use existing monitoring networks to estimate individual exposure to ambient air 

10 pollution; the inability to control for potential confounders such as other risk factors that 

11 affect birth outcomes (e.g., smoking); evaluating the exposure window (e.g., trimester) of 

12 importance; and limited evidence on the physiological mechanism of these effects (Ritz 

13 and Wilhelm, 2008; Slama et al., 2008). 

14 Overall, the evidence for an association between exposure to ambient O3 and 

15 reproductive and developmental outcomes is growing, yet remains relatively small. 

16 Recently, an international collaboration was formed to better understand the relationships 

17 between air pollution and adverse birth outcomes and to examine some of these 

18 methodological issues through standardized parallel analyses in datasets from different 

19 countries (Woodruff et al., 2010). Initial results from this collaboration have examined 

20 PM and birth weight (Parker et al., 2011); work on O3 has not yet been performed. 

21 Although early animal studies (Kavlock et al., 1980) found that exposure to O3 in the late 

22 gestation of pregnancy in rats led to some abnormal reproductive performances for 

23 neonates, to date human studies have reported inconsistent results for the association of 

24 ambient O3 concentrations and birth outcomes.

7.4.1 Effects on Sperm 

25 A limited amount of research has been conducted to examine the association between air 

26 pollution and male reproductive outcomes, specifically semen quality. To date, the 

27 epidemiologic studies have considered various exposure durations before semen 

28 collection that encompass either the entire period of spermatogenesis (i.e., 90 days) or 

29 key periods of sperm development that correspond to epididymal storage, development of 

30 sperm motility, and spermatogenesis. In an analysis conducted as part of the Teplice 

31 Program, 18-year-old men residing in the heavily polluted district of Teplice in the Czech 

32 Republic were found to be at greater risk of having abnormalities in sperm morphology 

33 and chromatin integrity than men of similar age residing in Prachatice, a less polluted 

34 district (Selevan et al., 2000; Sram et al., 1999). A follow-up longitudinal study 
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1 conducted on a subset of the same men from Teplice revealed associations between total 

2 episodic air pollution and abnormalities in sperm chromatin (Rubes et al., 2005). A 

3 limitation of these studies is that they did not identify specific pollutants or their 

4 concentrations.

5 More recent epidemiologic studies conducted in the U.S. have also reported associations 

6 between ambient air pollution and sperm quality for individual air pollutants, including 

7 O3 and PM2.5. In a repeated measures study in Los Angeles, CA, Sokol et al. (2006)

8 reported a reduction in average sperm concentration during three exposure windows 

9 (short-term exposures of 0-9, 10-14, and 70-90 days before semen collection, as well as 

10 long-term exposures of 0-90 days before semen collection) associated with high ambient 

11 levels of O3 in healthy sperm donors. This effect persisted under a joint additive model 

12 for O3, CO, NO2 and PM10. The authors did not detect a reduction in sperm count. Hansen 

13 et al. (2010) investigated the effect of exposure to O3 and PM2.5 (using the same exposure 

14 windows used by Sokol et al. (2006) on sperm quality in three southeastern counties 

15 (Wake County, NC; Shelby County, TN; Galveston County, TX). Outcomes included 

16 sperm concentration and count, morphology, DNA integrity and chromatin maturity. 

17 Overall, the authors found both protective and adverse effects, although some results 

18 suggested adverse effects on sperm concentration, count and morphology. 

19 The biological mechanisms linking ambient air pollution to decreased sperm quality have 

20 yet to be determined, though O3-induced oxidative stress, inflammatory reactions, and the 

21 induction of the formation of circulating toxic species have been suggested as possible 

22 mechanisms (see Section 5.3.8). Decremental effects on testicular morphology have been 

23 demonstrated in a toxicological study with histological evidence of O3-induced depletion 

24 of germ cells in testicular tissue and decreased seminiferous tubule epithelial layer. 

25 Jedlinska-Krakowska et al. (2006) demonstrated histopathological evidence of impaired 

26 spermatogenesis (round spermatids/ spermatocytes, giant spermatid cells, and focal 

27 epithelial desquamation with denudation to the basement membrane). The exposure 

28 protocol used five-month-old adult rats exposed to O3 as adults (long-term exposure, 

29 0.5 ppm, 5 h/day for 50 days). This degeneration could be rescued by vitamin E 

30 administration, indicating an antioxidant effect. Vitamin C administration had no effect at 

31 low doses of ascorbic acid and exacerbated the O3-dependent damage at high doses, as 

32 would be expected as vitamin C can be a radical generator instead of an antioxidant at 

33 higher doses. In summary, this study provided toxicological evidence of impaired 

34 spermatogenesis with O3 exposure that was rescued with certain antioxidant 

35 supplementation. 

36 Overall, there is limited epidemiologic evidence for an association with O3 concentration 

37 and decreased sperm concentration. A recent toxicological study provides limited 
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1 evidence for a possible biological mechanism (histopathology showing impaired 

2 spermatogenesis) for such an association.

7.4.2 Effects on Reproduction

3 Evidence suggests that exposure to air pollutants during pregnancy may be associated 

4 with adverse birth outcomes, which has been attributed to the increased sensitivity of the 

5 fetus due to physiologic immaturity. Gametes (i.e., ova and sperm) may be even more at6 risk, especially outside of the human body, as occurs with assisted reproduction. Smokers 

7 require twice the number of in vitro fertilization (IVF) attempts to conceive as non8 smokers (Feichtinger et al., 1997), suggesting that a preconception exposure can be 

9 harmful to pregnancy. A recent study used an established national-scale, log-normal 

10 kriging method to spatially estimate daily mean concentrations of criteria pollutants at 

11 addresses of women undergoing their first IVF cycle and at their IVF labs from 2000 to 

12 2007 in the northeastern U.S. (Legro et al., 2010). Increasing O3 concentration at the 

13 patient’s address during ovulation induction (short-term exposure, ~12 days) was 

14 significantly associated with an increased chance of live birth (OR = 1.13, [95% CI: 1.05, 

15 1.22] per 10 ppb increase), but with decreased odds of live birth when exposed from 

16 embryo transfer to live birth (long-term exposure, ~200 days) (OR = 0.79, [95% CI: 0.69, 

17 0.90] per 10 ppb increase). After controlling for NO2 in a copollutant model, however, O3

18 was no longer significantly associated with IVF failure. The results of this study suggest 

19 that short-term exposure to O3 during ovulation was beneficial (perhaps due to early 

20 conditioning to O3), whereas long-term exposure to O3 (e.g., during gestation) was 

21 detrimental, and reduced the likelihood of a live birth.

22 In most toxicological studies, reproductive success appears to be unaffected by O3

23 exposure. Nonetheless, one study has reported that 25% of the BALB/c mouse dams in 

24 the highest O3 exposure group (1.2 ppm, short-term exposure GD9-18) did not complete 

25 a successful pregnancy, a significant reduction (Sharkhuu et al., 2011). Ozone

26 administration (continuous 0.4, 0.8 or 1.2 ppm O3) to CD-1 mouse dams during the 

27 majority of pregnancy (short-term exposure, PD7-17, which excludes the 

28 pre-implantation period), led to no adverse effects on reproductive success (proportion of 

29 successful pregnancies, litter size, sex ratio, frequency of still birth, or neonatal mortality) 

30 (Bignami et al., 1994). There was a nearly statistically significant increase in pregnancy 

31 duration (0.8 and 1.2 ppm O3). Initially, dam body weight (0.8 and 1.2 ppm O3), water 

32 consumption (0.4, 0.8 and 1.2 ppm O3) and food consumption (0.4, 0.8 and 1.2 ppm O3) 

33 during pregnancy were decreased with O3 exposure but these deficits dissipated a week or 

34 two after the initial exposure (Bignami et al., 1994). The anorexigenic effect of O3

35 exposure on the pregnant dam appears to dissipate with time; the dams seem to adapt to 
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1 the O3 exposure. In males, data exist showing morphological evidence of altered 

2 spermatogenesis in O3 exposed animals (Jedlinska-Krakowska et al., 2006). Some 

3 evidence suggests that O3 may affect reproductive success when combined with other 

4 chemicals. Kavlock et al. (1979) showed that O3 acted synergistically with sodium 

5 salicylate to increase the rate of pup resorptions after midgestational exposure (1.0 ppm 

6 O3, short-term exposure, GD9-GD12). At low concentrations of O3 exposure, 

7 toxicological studies show reproductive effects to include a transient anorexigenic effect 

8 of O3 on gestational weight gain, and a synergistic effect of O3 on salicylate-induced pup 

9 resorptions; other fecundity, pregnancy- and gestation-related outcomes appear 

10 unaffected by O3 exposure. 

11 Collectively, there is very little epidemiologic evidence for the effect of short- or long12 term exposure to O3 on reproductive success, and the reproductive success in rats appears 

13 to be unaffected in toxicological studies of short-term exposure to O3. 

7.4.3 Birth Weight

14 With birth weight routinely collected in vital statistics and being a powerful predictor of 

15 infant mortality, it is the most studied outcome within air pollution-birth outcome 

16 research. Air pollution researchers have analyzed birth weight as a continuous variable 

17 and/or as a dichotomized variable in the form of LBW (<2,500 g [5 lbs, 8 oz]).

18 Birth weight is primarily determined by gestational age and intrauterine growth, but also 

19 depends on maternal, placental and fetal factors as well as on environmental influences. 

20 In both developed and developing countries, LBW is the most important predictor for 

21 neonatal mortality and is a significant determinant of postneonatal mortality and 

22 morbidity. Studies report that infants who are smallest at birth have a higher incidence of 

23 diseases and disabilities, which continue into adulthood (Hack and Fanaroff, 1999). 

24 The strongest evidence for an effect of O3 on birth weight comes from the Children’s 

25 Health Study conducted in southern California. In this study, Salam et al. (2005) report 

26 that maternal exposure to 24-h avg O3 concentrations averaged over the entire pregnancy 

27 was associated with reduced birth weight (39.3 g decrease [95% CI: -55.8, -22.8] in birth 

28 weight per 10 ppb and 8-h avg (19.2-g decrease [95% CI: -27.7, -10.7] in birth weight per 

29 10 ppb). This effect was stronger for concentrations averaged over the second and third 

30 trimesters. PM10, NO2 and CO concentrations averaged over the entire pregnancy were 

31 not statistically significantly associated with birth weight, although CO concentrations in 

32 the first trimester and PM10 concentrations in the third trimester were associated with a 

33 decrease in birth weight. Additionally, the authors observed a concentration-response 

34 relationship of birth weight with 24-h avg O3 concentrations averaged over the entire 
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1 pregnancy that was clearest above the 30-ppb level (see Figure 7-4). Relative to the 

2 lowest decile of 24-h avg O3, estimates for the next 5 lowest deciles were approximately 

3 -40 g to -50 g, with no clear trend and with 95% confidence bounds that included zero. 

4 The highest four deciles of O3 exposure showed an approximately linear decrease in birth 

5 weight, and all four 95% CIs excluded zero, and ranged from mean decreases of 

6 74 grams to decreases of 148 grams. 

Note: Deficits are plotted against the decile-group-specific median O3 exposure. Error bars represent 95% CIs. Indicator variables 

for each decile of O3 exposure (except the least-exposed group) were included in a mixed model.

Source: Salam et al. (2005).’

Figure 7-4 Birthweight deficit by decile of 24-h avg ozone concentration 

averaged over the entire pregnancy compared with the decile 

group with the lowest ozone exposure.

7 Several additional studies conducted in the U.S. and Canada also investigated the 

8 association between ambient O3 concentrations and birth weight and report some weak 

9 evidence for an association. Morello-Frosch et al. (2010) estimated ambient O3

10 concentrations throughout pregnancy and for each trimester in the neighborhoods of 

11 women who delivered term singleton births between 1996 and 2006 in California. A 

12 10-ppb increase in the O3 concentration averaged across the entire pregnancy was 

13 associated with a 5.7-g decrease (95% CI: -6.6, -4.9) in birth weight when exposures 

14 were calculated using monitors within 10 km of the maternal address at date of birth. 
Draft – Do Not Cite or Quote 7-51 June 2012

1 When the distance from the monitor was restricted to 3 km, the decrease in birth weight 

2 associated with a 10-ppb increase in O3 concentration was 8.9 g (95% CI: -10.6, -7.1). 

3 These results persisted in copollutant models and in models that stratified by trimester of 

4 exposure, SES, and race. Darrow et al. (2011b) did not observe an association with birth 

5 weight and O3 concentrations during two exposure periods of interest (i.e., the first month 

6 and last trimester), but did find an association with reduced birth weight when examining 

7 the cumulative air pollution concentration during the entire pregnancy period. 

8 Additionally, they observed effect modification by race and ethnicity, such that 

9 associations between birth weight and third-trimester O3 concentrations were 

10 significantly stronger in Hispanics and non-Hispanic African Americans than in non11 Hispanic whites. Chen et al. (2002) used 8-h avg O3 concentrations to create exposure 

12 variables based on average maternal exposure for each trimester. Ozone was not found to 

13 be related to birth weight in single-pollutant models, though the O3 effect during the third 

14 trimester was borderline statistically significant in a copollutant model with PM10.

15 Several studies found no association between ambient O3 concentrations and birth 

16 weight. Wilhelm and Ritz (2005) extended previous analyses of term LBW (Ritz et al., 

17 2000; Ritz and Yu, 1999) to include the period 1994-2000. The authors examined varying 

18 residential distances from monitoring stations to see if the distance affected risk 

19 estimation, exploring the possibility that effect attenuation may result from local pollutant 

20 heterogeneity inadequately captured by ambient monitors. As in their previous studies, 

21 the authors observed associations between elevated concentrations of CO and PM10 both 

22 early and late in pregnancy and risk of term LBW. After adjusting for CO and/or PM10

23 the authors did not observe associations between O3 and term LBW in any of their 

24 models. Brauer et al. (2008) evaluated the impacts of air pollution (CO, NO2, NO, O3, 

25 SO2, PM2.5, PM10) on birth weight for the period 1999-2002 using spatiotemporal 

26 residential exposure metrics by month of pregnancy in Vancouver, BC. Quantitative 

27 results were not presented for the association between O3 and LBW, though the authors 

28 observed associations that were largely protective. Dugandzic et al. (2006) examined the 

29 association between LBW and ambient levels of air pollutants by trimester of exposure 

30 among a cohort of term singleton births from 1988-2000. Though there was some 

31 indication of an association with SO2 and PM10, there were no effects for O3.

32 Similarly, studies conducted in Australia, Latin America, and Asia report limited 

33 evidence for an association between ambient O3 and measures of birth weight. In Sydney, 

34 Australia, Mannes et al. (2005) found that O3 concentrations in the second trimester of 

35 pregnancy had small adverse effects on birth weight (7.5-g decrease; [95% CI: -13.8, 1.2] 

36 per 10 ppb), although this effect disappeared when the analysis was limited to births with 

37 a maternal address within 5 km of a monitoring station (87.7-g increase; [95% CI: 10.5, 

38 164.9] per 10 ppb). Hansen et al. (2007) reported that trimester and monthly specific 
Draft – Do Not Cite or Quote 7-52 June 2012

1 exposures to all pollutants were not statistically significantly associated with a reduction 

2 in birth weight in Brisbane, Australia. In Sao Paulo, Brazil, Gouveia et al. (2004) found 

3 that O3 exhibited a small inverse relation with birth weight over the third trimester (6.0-g 

4 decrease; [95% CI: -30.8, 18.8] per 10 ppb). Lin et al. (2004b) reported a positive, though 

5 not statistically significant, exposure-response relationship for O3 during the entire 

6 pregnancy in a Taiwanese study. In a study performed in Korea, Ha et al. (2001) reported 

7 no O3 effect during the first trimester of pregnancy, but they found that during the third 

8 trimester of pregnancy O3 was associated with LBW (RR = 1.05 [95% CI: 1.02, 1.08] per 

9 10 ppb). 

Table 7-4 Brief Summary of Epidemiologic Studies of Birth Weight.

Study

Location

Sample Size Mean O3 (ppb) Exposure assessment

Effect Estimatea

(95% CI)

Salam et al. (2005) California, U.S.

(n = 3,901)

24-h avg:

27.3

8 h:

50.6

ZIP code level Entire pregnancy:

-39.3 g (-55.8, -22.8)

T1: -6.1 g (-16.8, 4.8)

T2: -20.0 g (-31.7, -8.4)

T3: -20.7 g (-32.1, -9.3)

Morello-Frosch et al. 

(2010)

California, U.S. 

(n = 3,545,177)

24-h avg:

23.5

Nearest Monitor 

(within 10, 5, 3 km)

Entire pregnancy:

-5.7 g (-6.6, -4.9)

T1: -2.1 g (-2.9, -1.4)

T2: -2.3 g (-3.1, -1.5)

T3: -1.3 g (-2.1, -0.6)

Darrow et al. (2011b) Atlanta, GA

(N=406,627)

8-h max:

44.8

Population-weighted 

spatial average

Entire pregnancy:

-12.3 g (-17.8, -6.8)

First 28 days

-0.5 g (-3.0, 2.1)

T3: -0.9g (-4.5, 2.8)

Chen et al. (2002) Northern Nevada, 

U.S.

(n = 36,305)

8-h:

27.2

County level Entire pregnancy:

20.9 g (6.3, 35.5)

T1: 23.4 g (-35.6 , 82.4)

T2: -19.4 g (-77.0, 38.2)

T3: 7.7 g (-50.9, 66.3)

Wilhelm and Ritz 

(2005)

Los Angeles County, 

CA

(n = 136,134)

1-h:

21.1-22.2

Varying distances from 

monitor

T1: NR

T3: NR

6 weeks before birth: NR

Brauer et al. (2008) Vancouver, BC, 

Canada

(n = 70,249)

24-h avg:

14

Nearest Monitor 

(within 10 km)

Inverse Distance 

Weighting (IDW)

Entire pregnancy: NR

First 30 days of pregnancy: NR

Last 30 days of pregnancy: NR

T1: NR

T3: NR

Dugandzic et al. 

(2006)

Nova Scotia, Canada

(n = 74,284)

24-h avg:

21

Nearest Monitor 

(within 25 km)

T1: 0.97 (0.81, 1.18)d

T2: 1.06 (0.87, 1.27)d

T3: 1.01 (0.83-1.24)d
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Study

Location

Sample Size Mean O3 (ppb) Exposure assessment

Effect Estimatea

(95% CI)

Mannes et al. (2005) Sydney, Australia

(n = 138,056)

1-h max:

31.6

Citywide avg and 

<5 km from monitor

T1: -0.9 g (-6.6, 4.8)

T2: -7.5 g (-13.8, 1.2)

T3: -4.5 g (-10.8, 1.8)

Last 30 days: 

-1.1 g (-5.6, 3.4)

Hansen et al. (2007) Brisbane, Australia

(n = 26,617)

8 h max:

26.7

Citywide avg T1: 2.8 g (-10.5, 16.0)

T2: 4.4 g (-11.4, 20.1)

T3: 11.3 g (-4.4, 27.1)

Gouveia et al. (2004) Sao Paulo, Brazil

(n = 179,460)

1-h max:

31.5

Citywide avg T1: -3.2 g (-25.6, 19)

T2: -0.2 g (-23.8, 23.4)

T3: -6.0 g (-30.8, -18.8)

Lin et al. (2004b) Kaohsiung and 

Taipei, Taiwan

(n = 92,288)

24-h avg: 15.86-

47.78

Nearest monitor 

(within 3 km)

Entire pregnancy:

1.13 (0.92, 1.38)c

T1: 1.02 (0.85, 1.22)c

T2: 0.93 (0.78, 1.12)c

T3: 1.05 (0.87, 1.26)c

Ha et al. (2001) Seoul, Korea

(n = 276,763)

8-h avg: 

22.4-23.3b

Citywide avg T1: 0.87 (0.81, 0.94)c

T3: 1.05 (1.02, 1.08)c

a

Change in birthweight per 10 ppb change in O3

b

Median

c

Odds ratios of LBW; Highest quartile of exposure compared to lowest quartile of exposure

d

Relative risk of LBW per 10 ppb change in O3

T1 = First Trimester, T2 = Second Trimester, T3 = Third Trimester

NR: No quantitative results reported

1 Table 7-4 provides a brief overview of the epidemiologic studies of birth weight. In 

2 summary, only the Children’s Health Study conducted in southern California (Salam et 

3 al., 2005) provides strong evidence for an effect of ambient O3 on birth weight. The study 

4 by Morello-Frosch et al. (2010), also conducted in California, provides support for the 

5 results of the Children’s Health Study. Additional studies, conducted in the U.S., Canada, 

6 Australia, Latin America, and Asia, provide limited and inconsistent evidence to support 

7 the effect reported in the Children’s Health Study. The toxicological literature on the 

8 effect of O3 on birth weight is sparse. In some studies, the reporting of birth weight may 

9 be avoided because birth weight can be confounded by decreased litter size resulting 

10 from an increased rate of pup resorption (aborted pups) in O3 exposed dams. In one 

11 toxicological study by Haro and Paz (1993), no differences in litter size were observed 

12 and decreased birth weight in pups from dams who were exposed to 1ppm O3 during 

13 pregnancy (short-term exposure, ~22 days) was reported. A second animal toxicology 

14 study recapitulated these finding with pregnant BALB/c mice that exposed to O3

15 (1.2 ppm, short-term exposure, GD9-18) producing pups with significantly decreased 

16 birth weights (Sharkhuu et al., 2011). 
Draft – Do Not Cite or Quote 7-54 June 2012

7.4.4 Preterm Birth

1 Preterm birth (PTB) is a syndrome (Romero et al., 2006) that is characterized by multiple 

2 etiologies. It is therefore unusual to be able to identify an exact cause for each PTB. In 

3 addition, PTB is not an adverse outcome in itself, but an important determinant of health 

4 status (i.e., neonatal morbidity and mortality). Although some overlap exists for common 

5 risk factors, different etiologic entities related to distinct risk factor profiles and leading 

6 to different neonatal and postneonatal complications are attributed to PTB and measures 

7 of fetal growth. Although both restricted fetal growth and PTB can result in LBW, 

8 prematurity does not have to result in LBW or growth restricted babies. 

9 A major issue in studying environmental exposures and PTB is selecting the relevant 

10 exposure period, since the biological mechanisms leading to PTB and the critical periods 

11 of vulnerability are poorly understood (Bobak, 2000). Short-term exposures proximate to 

12 the birth may be most relevant if exposure causes an acute effect. However, exposure 

13 occurring in early gestation might affect placentation, with results observable later in 

14 pregnancy, or cumulative exposure during pregnancy may be the most important 

15 determinant. The studies reviewed have dealt with this issue in different ways. Many 

16 have considered several exposure metrics based on different periods of exposure. Often 

17 the time periods used are the first month (or first trimester) of pregnancy and the 

18 last month (or 6 weeks) prior to delivery. Using a time interval prior to delivery 

19 introduces an additional problem since cases and controls are not in the same stage of 

20 development when they are compared. For example, a preterm infant delivered at 

21 36 weeks is a 32-week fetus 4 weeks prior to birth, while an infant born at term 

22 (40 weeks) is a 36-week fetus 4 weeks prior to birth. 

23 Recently, investigators have examined the association of PTB with both short-term 

24 (i.e., hours, days, or weeks) and long-term (i.e., months or years) exposure periods. Time25 series studies have been used to examine the association between air pollution 

26 concentrations during the days immediately preceding birth. An advantage of these time27 series studies is that this approach can remove the influence of covariates that vary across 

28 individuals over a short period of time. Retrospective cohort and case-control studies 

29 have been used to examine long-term exposure periods, often averaging air pollution 

30 concentrations over months or trimesters of pregnancy.

31 Studies of PTB fail to show consistency in pollutants and periods during pregnancy when 

32 an effect occurs. For example, while some studies find the strongest effects associated 

33 with exposures early in pregnancy, others report effects when the exposure is limited to 

34 the second or third trimester. However, the effect of air pollutant exposure during 

35 pregnancy on PTB has a biological basis. There is an expanding list of possible 
Draft – Do Not Cite or Quote 7-55 June 2012

1 mechanisms that may explain the association between O3 exposure and PTB (see 

2 Section 5.4.2.4).

3 Many studies of PTB compare exposure in quartiles, using the lowest quartile as the 

4 reference (or control) group. No studies use a truly unexposed control group. If exposure 

5 in the lowest quartile confers risk, than it may be difficult to demonstrate additional risk 

6 associated with a higher quartile. Thus negative studies must be interpreted with caution.

7 Preterm birth occurs both naturally (idiopathic PTB), and as a result of medical 

8 intervention (iatrogenic PTB). Ritz et al. (2007); (2000) excluded all births by Cesarean 

9 section to limit their studies to idiopathic PTB. No other studies attempted to distinguish 

10 the type of PTB, although air pollution exposure maybe associated with only one type. 

11 This is a source of potential effect misclassification. 

12 Generally, studies of air pollution and birth outcomes conducted in North America and 

13 the United Kingdom have not identified an association between PTB and maternal 

14 exposure to O3. Most recently, Darrow et al. (2009) used vital record data to construct a

15 retrospective cohort of 476,489 births occurring between 1994 and 2004 in 5 central 

16 counties of metropolitan Atlanta. Using a time-series approach, the authors examined 

17 aggregated daily counts of PTB in relation to ambient levels of CO, NO2, SO2, O3, PM10, 

18 PM2.5 and speciated PM measurements. This study investigated 3 gestational windows of 

19 short- and long-term exposure: the final week of gestation (short-term exposure), and the 

20 first month of gestation and the final 6 weeks of gestation (long-term exposure). The 

21 authors did not observe associations of PTB with O3 concentrations for any of the 

22 exposure periods.

23 A number of U.S. studies were conducted in southern California, and report somewhat 

24 inconsistent results. Ritz et al. (2000) evaluated the effect of air pollution (CO, NO2, O3, 

25 PM10) exposure during pregnancy on the occurrence of PTB in a cohort of 97,518 

26 neonates born in southern California between 1989 and 1993. The authors use both short27 and long-term exposure windows, averaging pollutant measures taken at the closest air28 monitoring station over distinct periods, such as 1, 2, 4, 6, 8, 12, and 26 weeks before 

29 birth and the whole pregnancy period. Additionally, they calculated average exposures 

30 for the first and second months of pregnancy. The authors found no consistent effects 

31 associated with O3 concentration over any of the pregnancy periods in single or 

32 multipollutant models. Wilhelm and Ritz (2005) extended previous analyses of PTB (Ritz 

33 et al., 2000; Ritz and Yu, 1999) in California to include 1994-2000. The authors 

34 examined varying residential distances from monitoring stations to see if the distance 

35 affected risk estimation, because effect attenuation may result from local pollutant 

36 heterogeneity inadequately captured by ambient monitors. The authors analyzed the 

37 association between long-term O3 exposure during varying periods of pregnancy and 
Draft – Do Not Cite or Quote 7-56 June 2012

1 PTB, finding a positive association between O3 levels in both the first trimester of 

2 pregnancy (RR = 1.23 [95% CI: 1.06, 1.42] per 10 ppb increase in 24-h avg O3) and the 

3 first month of pregnancy (results for first trimester exposure were similar, but slightly 

4 smaller, quantitative results not presented) in models containing all pollutants. No 

5 association was observed between O3 in the 6 weeks before birth and preterm delivery. 

6 Finally, Ritz et al. (2007) conducted a case-control survey nested within a birth cohort 

7 and assessed the extent to which residual confounding and exposure misclassification 

8 impacted air pollution effect estimates. The authors calculated mean long-term exposure 

9 levels for three gestational periods: the entire pregnancy, the first trimester, and the last 

10 6 weeks before delivery. Though positive associations were observed for CO and PM2.5, 

11 no consistent patterns of increase in the odds of PTB for O3 or NO2 were observed. 

12 A study conducted in Canada evaluated the impacts of air pollution (including CO, NO2, 

13 NO, O3, SO2, PM2.5, and PM10) on PTBs (1999-2002) using spatiotemporal residential 

14 exposure metrics by month of pregnancy (long-term exposure) in Vancouver, BC (Brauer 

15 et al., 2008). The authors did not observe consistent associations with any of the 

16 pregnancy average exposure metrics except for PM2.5 for PTB. The O3 associations were 

17 largely protective, and no quantitative results were presented for O3. Additionally, Lee et 

18 al. (2008c) used time-series techniques to investigate the associations of short-term 

19 exposure to O3 and PTB in London, England. In addition to exposure on the day of birth, 

20 cumulative exposure up to 1 week before birth was investigated. The risk of PTB did not 

21 increase with exposure to the levels of ambient air pollution experienced by this 

22 population. 

23 Conversely, studies conducted in Australia and China provide evidence for an association 

24 between ambient O3 and PTB. Hansen et al. (2006) reported that long-term exposure to 

25 O3 during the first trimester was associated with an increased risk of PTB (OR = 1.38, 

26 [95% CI: 1.14, 1.69] per 10 ppb increase). Although the test for trend was significant due 

27 to the strong effect in the highest quartile, there was not an obvious exposure-response 

28 pattern across the quartiles of O3 during the first trimester. The effect estimate was 

29 diminished and lost statistical significance when PM10 was included in the model 

30 (OR = 1.23, [95% CI: 0.97, 1.59] per 10 ppb increase). Maternal exposure to O3 during 

31 the 90 days prior to birth showed a weak, positive association with PTB (OR = 1.09, 

32 [95% CI: 0.85, 1.39] per 10 ppb increase). Jalaludin et al. (2007) found that O3 levels in 

33 the month and three months preceding birth had a statistically significant association with 

34 PTB. Ozone levels in the first trimester of pregnancy were associated with increased risks 

35 for PTBs (OR = 1.15 [95% CI: 1.05, 1.24] per 10 ppb increase in 1-h max O3

36 concentration), and remained a significant predictor of PTB in copollutant models (ORs 

37 between 1.07 and 1.10). Jiang et al. (2007) examined the effect of short- and long-term 

38 exposure to air pollution on PTB, including risk in relation to levels of pollutants for a 
Draft – Do Not Cite or Quote 7-57 June 2012

1 single day exposure window with lags from 0 to 6 days before birth. An increase of 

2 10 ppb of the 8-week avg of O3 corresponded to 9.47% (95% CI: 0.70, 18.7%) increase in 

3 PTBs. Increases in PTB were also observed for PM10, SO2, and NO2. The authors did not 

4 observe any significant effect of short-term exposure to outdoor air pollution on PTB 

5 among the 1-day time windows examined in the week before birth. 

6 Little data is available from toxicological studies; a study reported a nearly statistically 

7 significant increase in pregnancy duration (short-term exposure) in mice when exposed to 

8 0.8 or 1.2 ppm O3. This phenomenon was most likely due to the anorexigenic effect of 

9 relatively high O3 concentrations (Bignami et al., 1994).

10 Table 7-5 provides a brief overview of the epidemiologic studies of PTB. In summary, 

11 the evidence is consistent when examining short-term exposure to O3 during late 

12 pregnancy and reports no association with PTB. However when long-term exposure to O3

13 early in pregnancy is examined the results are inconsistent. Generally, studies conducted 

14 in the U.S., Canada, and England find no association with O3 and PTB, while studies 

15 conducted in Australia and China report an O3 effect on PTB. 
Draft – Do Not Cite or Quote 7-58 June 2012

Table 7-5 Brief summary of epidemiologic studies of PTB

Study

Location

Sample Size

Mean O3

(ppb)

Exposure 

assessment Effect Estimatea (95% CI)

Darrow et al. (2009) Atlanta, GA

(n = 476,489)

8-h max: 

44.1

Population-weighted 

spatial averages

Nearest Monitor 

(within 4 miles)

First month: 0.98 (0.97, 1.00)

Last week: 0.99 (0.98, 1.00)

Last 6 weeks: 1.00 (0.98, 1.02)

Ritz et al. (2000) California, U.S.

(n = 97,158)

8 h: 36.9 <2 mi of monitor First month: NR

Last 6 weeks: NR

Wilhelm and Ritz 

(2005)

Los Angeles, CA

(n = 106,483)

1 h: 21.1-

22.2

Varying distances to 

monitor

First month: 1.23 (1.06, 1.42)

T1: NR

T2: 1.38 (1.14, 1.66)

Last 6 weeks: NR

Ritz et al. (2007) Los Angeles, CA

(n = 58,316)

24-h avg: 

22.5

Nearest monitor to 

ZIP code

Entire pregnancy: NR

T1: 0.93 (0.82, 1.06)

Last 6 weeks: NR

Brauer et al. (2008) Vancouver, BC, 

Canada

(n = 70,249)

24-h avg:

14

Nearest Monitor 

(within 10 km)

Inverse Distance 

Weighting (IDW)

Entire pregnancy: NR

First 30 days of pregnancy: NR

Last 30 days of pregnancy: NR

T1: NR

T3: NR

Lee et al. (2008c) London, UK 24-h avg: NR 1 monitor Lag 0: 1.00 (1.00, 1.01)

Hansen et al. (2006) Brisbane, 

Australia

(n = 28,200)

8-h max:

26.7

Citywide avg T1: 1.39 (1.15, 1.70)

T3: 1.09 (0.88, 1.39)

Jalaludin et al. 

(2007)

Sydney, 

Australia

(n = 123,840)

1-h max:

30.9

Citywide avg and <5 

km from monitor

First month: 1.04 (0.95, 1.13)

T1: 1.15 (1.05, 1.24)

T3: 0.98 (0.89, 1.07)

Last month: 0.98 (0.88, 1.06)

Jiang et al. (2007) Shanghai, China

(n = 3,346 

preterm births)

8-h avg:

32.7

Citywide avg 4 wks before birth: 1.06 (1.00, 1.12)

6 wks before birth: 1.06 (0.99, 1.13)

8 wks before birth: 1.09 (1.01, 1.19)

L0: NR (results presented in figure)

L1: NR (results presented in figure)

L2: NR (results presented in figure)

L3: NR (results presented in figure)

L4: NR (results presented in figure)

L5: NR (results presented in figure)

L6: NR (results presented in figure)

a

Relative risk of PTB per 10 ppb change in O3.

T1 = First Trimester, T2 = Second Trimester, T3 = Third Trimester.

L0 = Lag 0, L1= Lag 1, L2 = Lag 2, L3 = Lag 3, L4 = Lag 4, L5 = Lag 5, L6 = Lag 6.

NR: No quantitative results reported.
Draft – Do Not Cite or Quote 7-59 June 2012

7.4.5 Fetal Growth 

1 Low birth weight has often been used as an outcome measure because it is easily 

2 available and accurately recorded on birth certificates. However, LBW may result from 

3 either short gestation, or inadequate growth in utero. Most of the studies investigating air 

4 pollution exposure and LBW limited their analyses to term infants to focus on inadequate 

5 growth. A number of studies were identified that specifically addressed growth restriction 

6 in utero by identifying infants who failed to meet specific growth standards. Usually 

7 these infants had birth weight less than the 10th percentile for gestational age, using an 

8 external standard. Many of these studies have been previously discussed, since they also 

9 examined other reproductive outcomes (i.e., LBW or PTB).

10 Fetal growth is influenced by maternal, placental, and fetal factors. The biological 

11 mechanisms by which air pollutants may influence the developing fetus remain largely 

12 unknown. Several mechanisms have been proposed, and are the same as those 

13 hypothesized for birth weight (see Section 5.4.2.4). Additionally, in animal toxicology 

14 studies, O3 causes transient anorexia in exposed pregnant dams. This may be one of many 

15 possible contributors to O3-dependent decreased fetal growth.

16 A limitation of environmental studies that use birth weight as a proxy measure of fetal 

17 growth is that patterns of fetal growth during pregnancy cannot be assessed. This is 

18 particularly important when investigating pollutant exposures during early pregnancy as 

19 birth weight is recorded many months after the exposure period. The insult of air 

20 pollution may have a transient effect on fetal growth, where growth is hindered at one 

21 point in time but catches up at a later point. For example, maternal smoking during 

22 pregnancy can alter the growth rate of individual body segments of the fetus at variable 

23 developmental stages, as the fetus experiences selective growth restriction and 

24 augmentation (Lampl and Jeanty, 2003).

25 The terms small-for-gestational-age (SGA), which is defined as a birth weight <10th 

26 percentile for gestational age (and often sex and/or race), and intrauterine growth 

27 retardation (IUGR) are often used interchangeably. However, this definition of SGA does 

28 have limitations. For example, using it for IUGR may overestimate the percentage of 

29 “growth-restricted” neonates as it is unlikely that 10% of neonates have growth 

30 restriction (Wollmann, 1998). On the other hand, when the 10th percentile is based on the 

31 distribution of live births at a population level, the percentage of SGA among PTB is 

32 most likely underestimated (Hutcheon and Platt, 2008). Nevertheless, SGA represents a 

33 statistical description of a small neonate, whereas the term IUGR is reserved for those 

34 with clinical evidence of abnormal growth. Thus all IUGR neonates will be SGA, but not 

35 all SGA neonates with be IUGR (Wollmann, 1998). In the following section the terms 

36 SGA and IUGR are referred to as each cited study used the terms.
Draft – Do Not Cite or Quote 7-60 June 2012

1 Over the past decade a number of studies examined various metrics of fetal growth 

2 restriction. Salam et al. (2005) assessed the effect of increasing O3 concentrations on 

3 IUGR in a population of infants born in California from 1975-1987 as part of the 

4 Children’s Health Study. The authors reported that maternal O3 exposures averaged over 

5 the entire pregnancy and during the third trimester were associated with increased risk of 

6 IUGR. A 10-ppb difference in 24-h maternal O3 exposure during the third trimester 

7 increased the risk of IUGR by 11% (95% CI: 0, 20%). Brauer et al. (2008) evaluated the 

8 impacts of air pollution (CO, NO2, NO, O3, SO2, PM2.5, PM10) on SGA (1999-2002) using 

9 spatiotemporal residential exposure metrics by month of pregnancy in Vancouver, BC. 

10 The O3 associations were largely protective (OR = 0.87, [95% CI: 0.81, 0.93] for a 

11 10 ppb increase in inverse distance weighted SGA), and no additional quantitative results 

12 were presented for O3. Liu et al. (2007b) examined the association between IUGR among 

13 singleton term live births and SO2, NO2, CO, O3, and PM2.5 in 3 Canadian cities for the 

14 period 1985-2000. No increase in the risk of IUGR in relation to exposure to O3 averaged 

15 over each month and trimester of pregnancy was noted. 

16 Three studies conducted in Australia provide evidence for an association between 

17 ambient O3 and fetal growth restriction. Hansen et al. (2007) examined SGA among 

18 singleton, full-term births in Brisbane, Australia in relation to ambient air pollution (bsp, 

19 PM10, NO2, O3) during pregnancy. They also examined head circumference and crown20 heel length in a subsample of term neonates. Trimester specific exposures to all pollutants 

21 were not statistically significantly associated with a reduction in head circumference or 

22 an increased risk of SGA. When monthly-specific exposures were examined, the authors 

23 observed an increased risk of SGA associated with exposure to O3 during month 4 

24 (OR = 1.11 [95% CI: 1.00, 1.24] per 10 ppb increase). In a subsequent study, Hansen et 

25 al. (2008) examined the possible associations between fetal ultrasonic measurements and 

26 ambient air pollution (PM10, O3, NO2, SO2) during early pregnancy. This study had two 

27 strengths: (1) fetal growth was assessed during pregnancy as opposed to at birth; and (2) 

28 there was little delay between exposures and fetal growth measurements, which reduces 

29 potential confounding and uses exposures that are concurrent with the observed growth 

30 pattern of the fetus. Fetal ultrasound biometric measurements were recorded for biparietal 

31 diameter (BPD), femur length, abdominal circumference, and head circumference. To 

32 further improve exposure assessment, the authors restricted the samples to include only 

33 scans from women for whom the centroid of their postcode was within 14 km of an air 

34 pollution monitoring site. Ozone during days 31-60 was associated with decreases in all 

35 of the fetal growth measurements, and a 1.78 mm reduction in abdomen circumference 

36 per 10 ppb increase in O3 concentration, though this effect did not persist in copollutant 

37 models. The change in ultrasound measurements associated with O3 during days 31-60 of 

38 gestation indicated that increasing O3 concentration decreased the magnitude of 

39 ultrasound measurements for women living within 2 km of the monitoring site. The 
Draft – Do Not Cite or Quote 7-61 June 2012

1 relationship decreased toward the null as the distance from the monitoring sites increased. 

2 When assessing effect modification due to SES, there was some evidence of effect 

3 modification for most of the associations, with the effects of air pollution stronger in the 

4 highest SES quartile. In the third study, Mannes et al. (2005) estimated the effects of 

5 pollutant (PM10, PM2.5, NO2, CO and O3) exposure in the first, second and third trimesters 

6 of pregnancy and risk of SGA in Sydney, Australia. Citywide average air pollutant 

7 concentrations in the last month, third trimester, and first trimester of pregnancy had no 

8 effect on SGA. Concentrations of O3 in the second trimester of pregnancy had small but 

9 adverse effects on SGA (OR = 1.10 [95% CI: 1.00, 1.14] per 10 ppb increment). This 

10 effect disappeared when the analysis was limited to births with a maternal address within 

11 5 km of a monitoring station (OR = 1.00 [95% CI: 0.60, 1.79] per 10 ppb increment).

12 Very little information from toxicological studies is available to address effects on fetal 

13 growth. However, there is evidence to suggest that prenatal (short-term) exposure to O3 

14 can affect postnatal growth. A few studies reported that mice or rats exposed 

15 developmentally (gestationally ± lactationally) to O3 had deficits in body weight gain in 

16 the postpartum period (Bignami et al., 1994; Haro and Paz, 1993; Kavlock et al., 1980). 

17 Table 7-6 provides a brief overview of the epidemiologic studies of fetal growth 

18 restriction. In summary, the evidence is inconsistent when examining exposure to O3 and 

19 fetal growth restriction. Similar to PTB, studies conducted in Australia have reported an 

20 effect of O3 on fetal growth, whereas studies conducted in other areas generally have not 

21 found such an effect. This may be due to the restriction of births to those within 2-14 km 

22 of a monitoring station, as was done in the Australian studies. 
Draft – Do Not Cite or Quote 7-62 June 2012

Table 7-6 Brief summary of epidemiologic studies of fetal growth.

Study Location

(Sample Size)

Mean O3 (ppb) Exposure 

assessment

Effect Estimatea (95% CI)

Salam et al. 

(2005)

California, U.S.

(n = 3901)

24-h avg:

27.3

8 h:

50.6

ZIP code level Entire pregnancy: 1.16 (1.00, 1.32)

T1:1.00 (0.94, 1.11)

T2: 1.06 (1.00, 1.12)

T3: 1.11 (1.00, 1.17)

Brauer et al. 

(2008)

Vancouver, BC, Canada

(n = 70,249)

24-h avg:

14

Nearest Monitor 

(within 10 km)

Inverse Distance 

Weighting (IDW)

Entire pregnancy: NR

First 30 days of pregnancy: NR

Last 30 days of pregnancy: NR

T1: NR

T3: NR

Liu et al. (2007b) Calgary, Edmonton, and 

Montreal, Canada

(n = 16,430)

24-h avg:

16.5

1-h max:

31.2

Census Subdivision 

avg

Entire pregnancy: NR (results 

presented in figure)

T1: NR (results presented in figure)

T2: NR (results presented in figure)

T3: NR (results presented in figure)

Hansen et al. 

(2007)

Brisbane, Australia

(n = 26,617)

8-h max:

26.7

Citywide avg T1: 1.01 (0.89, 1.15)

T2: 1.00 (0.86, 1.17)

T3: 0.83 (0.71, 0.97)

Hansen et al. 

(2008)

Brisbane, Australia

(n = 15,623)

8-h avg:

24.8

Within 2 km of monitor M1: -0.32 (-1.56, 0.91)b

M2: -0.58 (-1.97, 0.80)b

M3: 0.26 (-1.07, 1.59)b

M4: 0.11 (-0.98, 1.21)b

Mannes et al. 

(2005)

Sydney, Australia

(n = 138,056)

1-h max:

31.6

Citywide avg and 

<5 km from monitor

T1: 0.90 (0.48, 1.34)

T2: 1.00 (0.60, 1.79)

T3: 1.10 (0.66, 1.97)

Last 30 days of pregnancy: 1.10 (0.74, 

1.79)

a

Relative risk of fetal growth restriction per 10 ppb change in O3, unless otherwise noted.

b

Mean change in fetal ultrasonic measure of head circumference recorded between 13 and 26 weeks gestation for a 10-ppb 

increase in maternal exposure to O3 during early pregnancy

T1 = First Trimester, T2 = Second Trimester, T3 = Third Trimester

M1 = Month 1, M2 = Month 2, M3 = Month 3, M4 = Month 4

NR: No quantitative results reported

7.4.6 Postnatal Growth

1 Postnatal weight and height are routinely measured in children as indicators of growth 

2 and somatic changes. Toxicological studies often follow these endpoints to ascertain if a 

3 known exposure has an effect in the postnatal window, an effect which can be permanent. 

4 Time-pregnant BALB/c mice were exposed to O3 (0, 0.4, 0.8, or 1.2 ppm) GD9-18 

5 (short-term exposure) with parturition at GD20-21 (Sharkhuu et al., 2011). As the 

6 offspring aged, postnatal litter body weight continued to be significantly decreased in the 

7 highest concentration (1.2 ppm) O3 group at PND3 and PND7. When the pups were 
Draft – Do Not Cite or Quote 7-63 June 2012

1 weighed separately by sex at PND42, the males with the highest concentration of O3

2 exposure (1.2 ppm, GD9-18) had significant decrements in body weight (Sharkhuu et al., 

3 2011). 

4 Significant decrements in body weight at 4 weeks of age were reported in C57Bl/6 mice 

5 that were exposed to postnatal O3 (short-term exposure, PND2-28 exposure, 1 ppm O3, 

6 3 hours/day, 3 days/week) (Auten et al., 2012). Animals with co-exposure to in utero DE 

(short-term exposure, dam GD9-GD17; inhalation 0.5 or 2.0 mg/m3 7 O3; 4 h/day via 

8 inhalation; or oropharyngeal aspiration DEPs, 2×/week) + postnatal O3 (aforementioned 

9 short-term exposure) also had significantly reduced body weight.

7.4.7 Birth Defects 

10 Despite the growing body of literature evaluating the association between ambient air 

11 pollution and various adverse birth outcomes, relatively few studies have investigated the 

12 effect of temporal variations in ambient air pollution on birth defects. Heart defects and 

13 oral clefts have been the focus of the majority of these recent studies, given the higher 

14 prevalence than other birth defects and associated mortality. Mechanistically, air 

15 pollutants could be involved in the etiology of birth defects via a number of key events 

16 (see Section 5.4.2.4). 

17 Several studies have been conducted examining the relationship between O3 exposure 

18 during pregnancy and birth defects and reported a positive association with cardiac 

19 defects. The earliest of these studies was conducted in southern California (Ritz et al., 

20 2002). This study evaluated the effect of air pollution on the occurrence of cardiac birth 

21 defects in neonates and fetuses delivered in southern California in 1987-1993. Maternal 

22 exposure estimates were based on data from the fixed site closest to the mother’s ZIP 

23 code area. When using a case-control design where cases were matched to 10 randomly 

24 selected controls, results showed increased risks for aortic artery and valve defects 

25 (OR = 1.56 [95% CI: 1.16, 2.09] per 10 ppb O3), pulmonary artery and valve anomalies 

26 (OR = 1.34 [95% CI: 0.96, 1.87] per 10 ppb O3), and conotruncal defects (OR = 1.36 

27 [95% CI: 0.91, 2.03] per 10 ppb O3) in a dose-response manner with second-month O3

28 exposure. A study conducted in Texas (Gilboa et al., 2005) looked at a similar period of 

29 exposure but reported no association with most of the birth defects studied (O3

30 concentration was studied using quartiles with the lowest representing <18 ppb and the 

31 highest representing ≥ 31 ppb). The authors found slightly elevated odds ratios for 

32 pulmonary artery and valve defects. They also detected an inverse association between O3

33 exposure and isolated ventricular septal defects. Overall, this study provided some weak 

34 evidence that air pollution increases the risk of cardiac defects. Hansen et al. (2009)
Draft – Do Not Cite or Quote 7-64 June 2012

1 investigated the possible association between ambient air pollution concentrations 

2 averaged over weeks 3-8 of pregnancy and the risk of cardiac defects. When analyzing all 

3 births with exposure estimates for O3 from the nearest monitor there was no indication for 

4 an association with cardiac defects. There was also no adverse association when 

5 restricting the analyses to only include births where the mother resided within 12 km of a 

6 monitoring station. However, among births within 6 km of a monitor, a 10 ppb increase 

7 in O3 was associated with an increased risk of pulmonary artery and valve defects 

8 (OR = 8.76 [95% CI: 1.80, 56.55]). As indicated by the very wide credible intervals, 

9 there were very few cases in the sensitivity analyses for births within 6 km of a monitor, 

10 and this effect could be a result of type I errors. Dadvand et al. (2011) investigated the 

11 association between maternal exposure to ambient air pollution concentrations averaged 

12 over weeks 3-8 of pregnancy and the occurrence of cardiac birth defects in England.

13 Similar to Hansen et al. (2009), they found no associations with maternal exposure to O3

14 except for when the analysis was limited to those subjects residing within a 16 km 

15 distance of a monitoring station (OR for malformations of pulmonary and tricuspid 

16 valves=1.64 [95% CI: 1.04, 2.60] per 10 ppb increase in O3).

17 Despite the association between O3 and cardiac defects observed in the above studies, a 

18 recent study did not observe an increased risk of cardiac birth defects associated with 

19 ambient O3 concentrations. The study, conducted in Atlanta, GA, examined O3 exposure 

20 during weeks 3-7 of of pregnancy and reported no association with risk of cardiovascular 

21 malformations (Strickland et al., 2009). 

22 Several of these studies have also examined the relationship between O3 exposure during 

23 pregnancy and oral cleft defects. The study by Ritz et al. (2002) evaluated the effect of air 

24 pollution on the occurrence of orofacial birth defects and did not observe strong 

25 associations between ambient O3 concentration and orofacial defects. They did report an 

26 OR of 1.13 (95% CI: 0.90, 1.40) per 10 ppb during the second trimester for cleft lip with 

27 or without cleft palate. Similarly, Gilboa et al. (2005) reported an OR of 1.09 (95% CI: 

28 0.70, 1.69) for oral cleft defects when the fourth quartile was contrasted with the first 

29 quartile of exposure during 3-8 weeks of pregnancy. Hansen et al. (2009) reported no 

30 indication for an association with cleft defects and air pollution concentrations averaged 

31 over weeks 3-8 of pregnancy. Hwang and Jaakkola (2008) conducted a population-based 

32 case-control study to investigate exposure to ambient air pollution and the risk of cleft lip 

33 with or without cleft palate in Taiwan. The risk of cleft lip with or without cleft palate 

34 was increased in relation to O3 levels in the first gestational month (OR = 1.17 [95% CI: 

35 1.01, 1.36] per 10 ppb) and second gestational month (OR = 1.22 [95% CI: 1.03, 1.46] 

36 per 10 ppb), but was not related to any of the other pollutants. In three-pollutant models, 

37 the effect estimates for O3 exposure were stable for the four different combinations of 

38 pollutants and were all statistically significant. Marshall et al. (2010) compared estimated 
Draft – Do Not Cite or Quote 7-65 June 2012

1 exposure to ambient pollutants during early pregnancy (6 week period from 5 to 10 

2 weeks into the gestational period) among mothers of children with oral cleft defects to 

3 that among mothers of controls. The authors observed no consistent elevated associations 

4 between any of the air pollutants examined and cleft malformations, though there was a 

5 weak association between cases of cleft palate only and increasing O3 concentrations. 

6 This association increased when cases and controls were limited to those with residences 

7 within 10 km of the closest O3 monitor (OR = 2.2 [95% CI: 1.0, 4.9], comparing highest 

8 quartile [>33 ppb] to lowest quartile [<15 ppb]). 

9 A limited number of toxicological studies have examined birth defects in animals 

10 exposed gestationally to O3. Kavlock et al. (1979) exposed pregnant rats to O3 for precise 

11 periods during organogenesis. No significant teratogenic effects were found in rats 

12 exposed 8 h/day to concentrations of O3 varying from 0.44 to 1.97 ppm during early 

13 (days 6-9), mid (days 9-12), or late (days 17 to 20) gestation, or the entire period of 

14 organogenesis (days 6-15) (short-term exposures). Earlier research found eyelid 

15 malformation following gestational and postnatal exposure to 0.2 ppm O3 (Veninga, 

16 1967). 

17 Table 7-7 provides a brief overview of the epidemiologic studies of birth defects. These 

18 studies have focused on cardiac and oral cleft defects, and the results from these studies 

19 are not entirely consistent. This inconsistency could be due to the absence of true 

20 associations between O3 and risks of cardiovascular malformations and oral cleft defects; 

21 it could also be due to differences in populations, pollution levels, outcome definitions, or 

22 analytical approaches. The lack of consistency of associations between O3 and 

23 cardiovascular malformations or oral cleft defects might be due to issues relating to 

24 statistical power or measurement error. A recent meta-analysis of air pollution and 

25 congenital anomalies concluded that there was no statistically significant increase in risk 

26 of congenital anomalies and O3 (Vrijheid et al., 2011). These authors note that 

27 heterogeneity in the results of these studies may be due to inherent differences in study 

28 location, study design, and/or analytic methods, and comment that these studies have not 

29 employed some recent advances in exposure assessment used in other areas of air 

30 pollution research that may help refine or reduce this heterogeneity.
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Table 7-7 Brief summary of epidemiologic studies of birth defects

Study

Outcomes 

Examined

Location

(Sample Size) Mean O3 (ppb)

Exposure 

Assessment

Exposure 

Window

Ritz et al. (2002) Cardiac and Cleft 

Defects

Southern California

(n = 3,549 cases; 

10,649 controls)

24-h avg:

NR

Nearest Monitor 

(within 10 mi)

Month 1,2,3

Trimester 2,3

3-mo period prior to 

conception

Gilboa et al. (2005) Cardiac and Cleft 

Defects

7 Counties in TX

(n = 5,338 cases; 

4,580 controls)

24-h avg:

NR

Nearest Monitor Weeks 3-8 of 

gestation

Hwang and Jaakkola 

(2008)

Oral Cleft Defects Taiwan

(n = 653 cases; 

6,530 controls)

24-h avg:

27.31

Inverse Distance 

Weighting (IDW)

Months 1,2,3

Strickland et al. 

(2009)

Cardiac Defects Atlanta, GA

(n = 3,338 cases)

8-h max:

39.8-43.3

Weighted citywide 

avg

Weeks 3-7 of 

gestation

Hansen et al. (2009) Cardiac and Cleft 

Defects

Brisbane, Australia

(n = 150,308 births)

8-h max:

25.8

Nearest Monitor Weeks 3-8 of 

gestation

Marshall et al. (2010) Oral Cleft Defects New Jersey

(n = 717 cases; 

12,925 controls)

24-h avg:

25

Nearest Monitor 

(within 40 km)

Weeks 5-10 of 

gestation

Dadvand et al. 

(2011)

Cardiac Defects Northeast England

(n = 2,140 cases; 

14,256 controls)

24-h avg:

18.8

Nearest Monitor Weeks 3-8 of 

gestation`1

7.4.8 Developmental Respiratory Effects

1 The issue of prenatal exposure has assumed increasing importance since ambient air 

2 pollution exposures of pregnant women have been shown to lead to adverse pregnancy 

3 outcomes, as well as to respiratory morbidity and mortality in the first year of life. 

4 Growth and development of the respiratory system take place mainly during the prenatal 

5 and early postnatal periods. This early developmental phase is thought to be very 

6 important in determining long-term lung growth. Studies have recently examined this 

7 emerging issue. Several studies were included in Section 7.2.1 and Section 7.2.3, and are 

8 included here because they reported both prenatal and post-natal exposure periods. 

9 Mortimer et al. (2008a); (2008b) examined the association of prenatal and lifetime 

10 exposures to air pollutants with pulmonary function and allergen sensitization in a subset 

11 of asthmatic children (ages 6-11) included in the Fresno Asthmatic Children’s 

12 Environment Study (FACES). Monthly means of pollutant levels for the years 1989-2000 

13 were created and averaged separately across several important developmental time14 periods, including the entire pregnancy, each trimester, the first 3 years of life, the first 

15 6 years of life, and the entire lifetime. The 8-h avg O3 concentrations were approximately 
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1 50 ppb for each of the exposure metrics (estimated from figure). In the first analysis 

2 (Mortimer et al., 2008a), negative effects on pulmonary function were found for exposure 

3 to PM10, NO2, and CO during key neonatal and early life developmental periods. The 

4 authors did not find a negative effect of exposure to O3 among this cohort. In the second 

5 analysis (Mortimer et al., 2008b), sensitization to at least one allergen was associated, in 

6 general, with higher levels of CO and PM10 during the entire pregnancy and second 

7 trimester and higher PM10 during the first 2 years of life. Lower exposure to O3 during the 

8 entire pregnancy or second trimester was associated with an increased risk of allergen 

9 sensitization. Although the pollutant metrics across time periods are correlated, the 

10 strongest associations with the outcomes were observed for prenatal exposures. Though it 

11 may be difficult to disentangle the effect of prenatal and postnatal exposures, the models 

12 from this group of studies suggest that each time period of exposure may contribute 

13 independently to different dimensions of school-aged children’s pulmonary function. For 

14 4 of the 8 pulmonary-function measures (FVC, FEV1, PEF, FEF25-75), prenatal exposures 

15 were more influential on pulmonary function than early-lifetime metrics, while, in 

16 contrast, the ratio of measures (FEV1/FVC and FEF25-75/FVC) were most influenced by 

17 postnatal exposures. When lifetime metrics were considered alone, or in combination 

18 with the prenatal metrics, the lifetime measures were not associated with any of the 

19 outcomes, suggesting the timing of the exposure may be more important than the overall 

20 dose and prenatal exposures are not just markers for lifetime or current exposures. 

21 Clark et al. (2010) investigated the effect of exposure to ambient air pollution in utero 

22 and during the first year of life on risk of subsequent asthma diagnosis (incident asthma 

23 diagnosis up to age 3-4) in a population-based nested case-control study. Air pollution 

24 exposure for each subject based on their residential address history was estimated using 

25 regulatory monitoring data, land use regression modeling, and proximity to stationary 

26 pollution sources. An average exposure was calculated for the duration of pregnancy 

27 (~15 ppb) and the first year of life (~14 ppb). In contrast to the Mortimer et al. (2008a); 

28 (2008b) studies, the effect estimates for first-year exposure were generally larger than for 

29 in utero exposures. However, similar to the Mortimer et al. studies, the observed 

30 associations with O3 were largely protective. Because of the relatively high correlation 

31 between in utero and first-year exposures for many pollutants, it was difficult to discern 

32 the relative importance of the individual exposure periods. 

33 Latzin et al. (2009) examined whether prenatal exposure to air pollution was associated 

34 with lung function changes in the newborn. Tidal breathing, lung volume, ventilation 

35 inhomogeneity and eNO were measured in 241 unsedated, sleeping neonates (age= 

36 5 weeks). The median of the 24-h avg O3 concentrations averaged across the post-natal 

37 period was ~44 ppb. Consistent with the previous studies, no association was found for 

38 prenatal exposure to O3 and lung function.
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1 The new toxicological literature since the 2006 O3 AQCD, covering respiratory changes 

2 related to developmental O3 exposure, reports ultrastructural changes in bronchiole 

3 development, alterations in placental and pup cytokines, and increased pup airway hyper4 reactivity. These studies are detailed below. 

5 Fetal rat lung bronchiole development is triphasic, comprised of the glandular phase 

6 (measured at GD18), the canalicular phase (GD20), and the saccular phase (GD21). The 

7 ultrastructural lung development in fetuses of pregnant rats exposed to 1-ppm O3 (12 

8 h/day, out to either GD18, GD20 or GD21) was examined by electron microscopy during 

9 these three phases. In the glandular phase, bronchiolar columnar epithelial cells in fetuses 

10 of dams exposed to O3 had cytoplasmic damage and swollen mitochondria. Bronchial 

11 epithelium at the canalicular phase in O3 exposed pups had delayed maturation in 

12 differentiation, i.e., glycogen abundance in secretory cells had not diminished as it should 

13 with this phase of development. Congruent with this finding, delayed maturation of 

14 tracheal epithelium following early neonatal O3 exposure (1 ppm, 4-5 h/day for first week 

15 of life) in lambs has been previously reported (Mariassy et al., 1990; Mariassy et al., 

16 1989). Also at the canalicular phase, atypical cells were seen in the bronchiolar lumen of 

17 O3-exposed rat fetuses. Finally, in the saccular phase, mitochondrial degradation was 

18 present in the non-ciliated bronchiolar cells of rats exposed in utero to O3. In conclusion, 

19 O3 exposure of pregnant rats produced ultra-structural damage to near-term fetal 

20 bronchiolar epithelium (López et al., 2008). 

21 Exposure of laboratory animals to multiple airborne pollutants can differentially affect 

22 pup physiology. One study showed that exposure of C57BL/6 mouse dams to 0.48 mg 

23 PM intratracheally twice weekly for 3 weeks during pregnancy augmented O3-induced 

24 airway hyper-reactivity in juvenile offspring. Maternal PM exposure also significantly 

25 increased placental cytokines above vehicle-instilled controls. Pup postnatal O3 exposure 

26 (1 ppm 3 h/day, every other day, thrice weekly for 4 weeks) induced significantly 

27 increased cytokine levels (IL-1β, TNF-α, KC, and IL-6) in whole lung versus postnatal 

28 air exposed groups; this was further exacerbated with gestational PM exposure (Auten et 

29 al., 2009). In further studies by the same laboratory, O3-induced AHR was studied in 

30 rodent offspring after dam gestational exposure to inhaled diesel exhaust (Auten et al., 

2012). Pregnant C57Bl/6 mice were exposed to diesel exhaust GD9-17 (0.5 or 2.0 mg/m3 31

32 O3, 4h/day) via inhalation or in a separate set of animals via oropharyngeal aspiration of 

33 freshly generated DEPs (2×/week). Postnatally, the offspring were exposed to O3 starting 

34 at PND2 (1 ppm O3, 3 hours/day, 3 days/week for 4 weeks). Juvenile mice were then 

35 subjected to measurements of pulmonary mechanisms (at 4 weeks of age and then at 8 

36 weeks of age). Increased inflammation of the placenta and lungs of DE exposed fetuses 

37 was reported at GD18. In animals with postnatal O3 exposure alone, elevated 

38 inflammation was seen with significant increased levels of BAL cytokines; these O3-
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1 related elevated levels were significantly exacerbated with prenatal DE exposure 

2 (DE+O3). At PND28, DE+O3 exposed offspring had significant impairment of alveolar 

3 development as measured with secondary alveolar crest development, a finding that was 

4 absent in all other exposure groups (O3 alone, DE alone). Postnatal O3 exposure induced 

5 AHR in methacholine challenged animals at 4 weeks of age and was exacerbated with the 

6 higher dose of DE exposure (DE+O3). At 8 weeks of age, O3 exposed pups had persistent 

7 AHR (+/-DE) that was significantly augmented in DE+O3 pups. In summary, gestational 

8 DE exposure induced an inflammatory response which, when combined with postnatal O3

9 exposure impaired alveolar development, and caused an exacerbated and longer-lasting 

10 O3-induced AHR in offspring.

11 A series of experiments using infant rhesus monkeys repeatedly exposed to 0.5 ppm O3

12 starting at one-month of age have examined the effect of O3 alone or in combination with 

13 an inhaled allergen on morphology and lung function (Plopper et al., 2007). Exposure to 

14 O3 alone or allergen alone produced small but not statistically significant changes in 

15 baseline airway resistance and airway responsiveness, but the combined exposure to both 

16 O3 + antigen produced statistically significant and greater than additive changes in both 

17 functional measurements. Additionally, cellular changes and significant structural 

18 changes in the respiratory tract have been observed in infant rhesus monkeys exposed to 

19 O3 (Fanucchi et al., 2006). A more detailed description of these studies can be found in 

20 Section 7.2.3 (Pulmonary Structure and Function), with mechanistic information found in 

21 Section 5.4.2.4.

22 Lung immunological response in O3 exposed pups was followed by analyzing BAL and 

23 lung tissue. Sprague Dawley (SD) pups were exposed to a single 3h exposure of air or O3

24 (0.6 ppm) on PND 13 (Han et al., 2011). Bronchoalveolar lavage (BAL) was performed 

25 10 hours after the end of O3 exposure. BALF polymorphonuclear leukocytes (PMNs) and 

26 total BALF protein were significantly elevated in O3 exposed pups. Lung tissue from O3

27 exposed pups had significant elevations of manganese superoxide dismutase (SOD) 

28 protein and significant decrements of extra-cellular SOD protein.

29 Various immunological outcomes were followed in offspring after their pregnant dams 

30 (BALB/c mice) were exposed gestationally to O3 (0, 0.4, 0.8, or 1.2 ppm, GD9-18) 

31 (Sharkhuu et al., 2011). Delayed type hypersensitivity (DTH) was initiated with initial 

32 BSA injection at 6 weeks of age and then challenge 7 days later. The normal edematous 

33 response of the exposed footpad (thickness after BSA injection) was recorded as an 

34 indicator of DTH. In female offspring, normal footpad swelling with BSA injection that 

35 was seen in air exposed animals was significantly attenuated with O3 exposure (0.8 and 

36 1.2 ppm O3), implying immune suppression of O3 exposure specifically in DTH. Humoral 

37 immunity was measured with the sheep red blood cell (SRBC) response. Animals 
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1 received primary immunization with SRBC and then blood was drawn for SRBC IgM 

2 measurement. A SRBC booster was given 2 weeks later with blood collected 5 days after 

3 booster for IgG measurement. Maternal O3 exposure had no effect on humoral immunity 

4 in the offspring as measured by IgG and IgM titers after SRBC primary and booster 

5 immunizations (Sharkhuu et al., 2011). 

6 Toxicity assessment and allergen sensitization was also assessed in these O3 exposed 

7 offspring. At PND42, animals were euthanized for analysis of immune and inflammatory 

8 markers (immune proteins, inflammatory cells, T-cell populations in the spleen). A subset 

9 of the animals was intra-nasally instilled or sensitized with ovalbumin on either PND2 

10 and 3 or PND42 and 43. All animals were challenged with OVA on PND54, 55, and 56. 

11 One day after final OVA challenge, lung function, lung inflammation and immune 

12 response were determined. Offspring of O3 exposed dams that were initially sensitized at 

13 PDN3 (early) or PND42 (late) were tested to determine the level of allergic sensitization 

14 or asthma-like inflammation after OVA challenge. Female offspring sensitized early in 

15 life developed significant eosinophilia (1.2 ppm O3) and elevated serum OVA-specific 

16 IgE (1.2 ppm O3), which is a marker of airway allergic inflammation. The females that 

17 were sensitized early also had significant decrements in BALF total cells, macrophages, 

18 and lymphocytes (1.2 ppm O3). Offspring that were sensitized later (PND42) in life did 

19 not develop the aforementioned changes in BALF, but these animals did develop modest, 

20 albeit significant neutropenia (0.8 and 1.2 ppm O3) (Sharkhuu et al., 2011).

21 BALF cytology in non-sensitized animals was followed. BALF of offspring born to dams 

22 exposed to O3 was relatively unaffected (cytokines, inflammatory cell numbers/types) as 

23 were splenic T-cell subpopulations. LDH was significantly elevated in BALF of females 

24 whose mothers were exposed to 1.2 ppm during pregnancy (Sharkhuu et al., 2011). In 

25 summary, the females born to mothers exposed to O3 developed modest 

26 immunocompromise. Males were unaffected (Sharkhuu et al., 2011). 

27 Overall, animal toxicological studies have reported ultrastructural changes in bronchiole 

28 development, alterations in placental and pup cytokines, and increased pup airway hyper29 reactivity related to exposure to O3 during the developmental period. Epidemiologic 

30 studies have found no association between prenatal exposure to O3 and growth and 

31 development of the respiratory system. Fetal origins of disease have received a lot of 

32 attention recently, thus additional research to further explore the inconsistencies between 

33 these two lines of evidence is warranted.
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7.4.9 Developmental Central Nervous System Effects

1 The following sections describe the results of toxicological studies of O3 and 

2 developmental central nervous system effects. No epidemiologic studies of this 

3 association have been published.

7.4.9.1 Laterality

4 Two reports of laterality changes in mice developmentally exposed to O3 have been 

5 reported in the literature. Mice developmentally exposed to 0.6 ppm O3 (6 days before 

6 breeding to weaning at PND21) showed a turning preference (left turns) distinct from air 

7 exposed controls (clockwise turns) (Dell'Omo et al., 1995); in previous studies this 

8 behavior in mice has been found to correlate with specific structural asymmetries of the 

9 hippocampal mossy fiber projections (Schöpke et al., 1991). The 2006 O3 AQCD 

10 evidence for the effect of O3 on laterality or handedness demonstrated that rats exposed to 

11 O3 during fetal and neonatal life showed limited, gender-specific changes in handedness 

12 after exposure to the intermediate concentration of O3 (only seen in female mice exposed 

13 to 0.6 ppm O3, and not in males at 0.6 ppm or in either sex of 0.3 or 0.9 ppm O3 with 

14 exposure from 6 days before breeding to PND26) (Petruzzi et al., 1999). 

7.4.9.2 Brain Morphology and Neurochemical Changes

15 The nucleus tractus solitarius (NTS), a medullary area of respiratory control, of adult 

16 animals exposed prenatally to 0.5 ppm O3 (12h/day, ED5-eD20) had significantly less 

17 tyrosine hydroxylase staining versus control (Boussouar et al., 2009). Tyrosine 

18 hydroxylase is the rate-limiting enzyme for dopamine synthesis and serves as a precursor 

19 for catecholamine synthesis; thus, decreased staining is used as a marker of dopaminergic 

20 or catecholaminergic cell or activity loss in these regions and thus functions in neuronal 

21 plasticity. After physical restraint stress, control animals respond at the histological level 

22 with Fos activation, a marker of neuronal activity, and tyrosine hydroxylase activation in 

23 the NTS, a response which is absent or attenuated in adult animals exposed prenatally to 

24 0.5 ppm O3 (Boussouar et al., 2009) when compared to control air exposed animals who 

25 also were restrained. The O3-exposed offspring in this study were cross-fostered to 

26 control air exposed dams to avoid O3-dependent dam related neonatal effects on offspring 

27 outcomes (i.e., dam behavioral or lactational contributions to pup outcomes) (Boussouar 

28 et al., 2009). 
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1 Developmental exposure to 0.3 or 0.6 ppm O3 prior to mating pair formation through 

2 GD17 induced significant increased levels of BDNF in the striatum of adult (PND140) 

3 O3 exposed offspring as compared to control air exposed animals; these O3-exposed 

4 animals also had significantly decreased level of NGF in the hippocampus versus control 

5 (Santucci et al., 2006). 

6 Changes in the pup cerebellum with prenatal 1 ppm O3 exposure include altered 

7 morphology (Romero-Velazquez et al., 2002; Rivas-Manzano and Paz, 1999), decreased 

8 total area (Romero-Velazquez et al., 2002), decreased number of Purkinje cells (Romero9 Velazquez et al., 2002), and altered monoamine neurotransmitter content with the 

10 catecholamine system affected and the indoleamine system unaffected by O3 (Gonzalez11 Pina et al., 2008).

7.4.9.3 Neurobehavioral Outcomes

12 O3 administration to dams during pregnancy with or without early neonatal exposure has 

13 been shown to contribute to multiple neurobehavioral outcomes in offspring that are 

14 described in further detail below.

15 O3 administration (0.4, 0.8 or 1.2 ppm O3) during the majority of pregnancy (PD7-17) of 

16 CD-1 mice did not affect pup behavioral outcomes including early behavioral ultrasonic 

17 vocalizations and more permanent later measurements (PND60 or 61) including pup 

18 activity, habituation and exploration and d-amphetamine-induced hyperactivity (Bignami 

19 et al., 1994); these pups were all cross-fostered or reared on non- O3 exposed dams. 

20 Testing for aggressive behavior in mice continuously exposed to O3 (0.3 or 0.6 ppm from 

21 30 days prior to mating to GD17) revealed that mice had significantly increased 

22 defensive/ submissive behavior (increased freezing posturing on the first day only of a 

23 multiple-day exam) versus air exposed controls (Santucci et al., 2006). Similarly, 

24 continuous exposure of adult animals to O3 induced significant increases in fear behavior 

25 and decreased aggression as measured by significantly decreased freezing behavior 

26 (Petruzzi et al., 1995). 

27 Developmentally exposed animals also had significantly decreased amount of time spent 

28 nose sniffing other mice (Santucci et al., 2006); this social behavior deficit, decreased 

29 sniffing time, was not found in an earlier study with similar exposures (Petruzzi et al., 

30 1995), but sniffing of specific body areas was measured in Santucci et al. (2006) and total 

31 number of sniffs of the entire body was measured in Petruzzi et al. (1995). The two 

32 toxicology studies exploring social behavior (sniffing) employ different study designs 

33 and find opposite effects in animals exposed to O3.
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7.4.9.4 Sleep Aberrations after Developmental Ozone Exposure

1 The effect of gestational O3 exposure (1 ppm O3 daily for 12h/day, during dark period for 

2 the entire pregnancy) on sleep patterns in rat offspring was followed using 24 h 

3 polysomnographic recordings at 30, 60 and 90 days of age (Haro and Paz, 1993). 

4 Ozone-exposed pups manifested with inverted sleep-wake patterns or circadian rhythm 

5 phase-shift. Rat vigilance was characterized in wakefulness, slow wave sleep (SWS), and 

6 paradoxical sleep (PS) using previously characterized criteria. The O3 exposed offspring 

7 spent longer time in the wakefulness state during the light period, more time in SWS 

8 during the period of darkness, and showed significant decrements in PS. Chronic O3

9 inhalation significantly decreased the duration of PS during both the light and dark 

10 periods (Haro and Paz, 1993). These effects were consistent at all time periods measured 

11 (30, 60 and 90 days of age). These sleep effects reported after developmental exposures 

12 expand upon the existing literature on sleep aberrations in adult animals exposed to O3

13 [rodents: (Paz and Huitron-Resendiz, 1996; Arito et al., 1992); and cats: (Paz and Bazan14 Perkins, 1992)]. A role for inhibition of cyclooxygenase-2 and the interleukins and 

15 prostaglandins in the O3-dependent sleep changes potentially exists with evidence from a 

16 publication on indomethacin pretreatment attenuating O3-induced sleep aberrations in 

17 adult male animals (Rubio and Paz, 2003). 

7.4.10 Early Life Mortality

18 Infants may be particularly at risk for the effects of air pollution. Within the first year of 

19 life, infants develop rapidly; therefore their sensitivity may change within weeks or 

20 months. During the neonatal and post-neonatal periods, the developing lung is highly 

21 sensitive to environmental toxicants. The lung is not well developed at birth, with 80% of 

22 alveoli being formed postnatally. An important question regarding the association 

23 between O3 and infant mortality is the critical window of exposure during development 

24 for which infants are at risk. Several age intervals have been explored: neonatal 

25 (<1 month); postneonatal (1 month to 1 year); and an overall interval for infants that 

26 includes both the neonatal and postneonatal periods (<1 year). Within these various age 

27 categories, multiple causes of deaths have been investigated, particularly total deaths and 

28 respiratory-related deaths. The studies reflect a variety of study designs, exposure 

29 periods, regions, and adjustment for confounders. As discussed below, a handful of 

30 studies have examined the effect of ambient air pollution on neonatal and postneonatal 

31 mortality, with the former the least studied. These studies varied somewhat with regard to 

32 the outcomes and exposure periods examined and study designs employed.
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7.4.10.1 Stillbirth 

1 Pereira et al. (1998) investigated the association among daily counts of intrauterine 

2 mortality (over 28 weeks of gestation) and air pollutant concentrations in Sao Paulo, 

3 Brazil from 1991 through 1992. The association was strong for NO2, but lesser for SO2

4 and CO. These associations exhibited a short lag time, less than 5 days. No significant 

5 association was detected between short-term O3 exposure and intrauterine mortality.

7.4.10.2 Infant Mortality, Less than 1 Year 

6 Ritz et al. (2006) linked birth and death certificates for infants who died between 1989 

7 and 2000 to evaluate the influence of outdoor air pollution on infant death in the South 

8 Coast Air Basin of California. The authors examined short- and long-term exposure 

9 periods 2 weeks, 1 month, 2 months, and 6 months before each case subject’s death and 

10 reported no association between ambient levels of O3 and infant mortality. Similarly, 

11 Diaz et al. (2004) analyzed the effects of extreme temperatures and short-term exposure 

12 to air pollutants on daily mortality in children less than 1 year of age in Madrid, Spain, 

13 from 1986 to 1997 and observed no statistically significant association between mortality 

14 and O3 concentrations. Hajat et al. (2007) analyzed time-series data of daily infant 

15 mortality counts in 10 major cities in the UK to quantify any associations with short-term 

16 changes in air pollution. When the results from the 10 cities were combined there was no 

17 relationship between O3 and infant mortality, even after restricting the analysis to just the 

18 summer months. 

19 Conversely, a time-series study of infant mortality conducted in the southwestern part of 

20 Mexico City in the years 1993-1995 found that infant mortality was associated with 

21 short-term exposure to NO2 and O3 3-5 days before death, but not as consistently as with 

22 PM. A 10-ppb increase in 24-h avg O3 was associated with a 2.78% increase (95% CI: 

23 0.29, 5.26%) in infant mortality (lag 3) (Loomis et al., 1999). This increase was 

24 attenuated, although still positive when evaluated in a two-pollutant model with PM2.5. 

25 One-hour max concentrations of O3 exceeded prevailing Mexican and international 

26 standards nearly every day. 

7.4.10.3 Neonatal Mortality, Less than 1 Month

27 Several studies have evaluated ambient O3 concentrations and neonatal mortality and 

28 observed no association. Ritz et al. (2006) linked birth and death certificates for infants 

29 who died between 1989 and 2000 to evaluate the influence of outdoor air pollution on 
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1 infant death in the South Coast Air Basin of California. The authors examined short- and 

2 long-term exposure periods 2 weeks, 1 month, 2 months, and 6 months before each case 

3 subject’s death and reported no association between ambient levels of O3 and neonatal 

4 mortality. Hajat et al. (2007) analyzed time-series data of daily infant mortality counts in 

5 10 major cities in the UK to quantify any associations with short-term changes in air 

6 pollution. When the results from the 10 cities were combined there was no relationship 

7 between O3 and neonatal mortality, even after restricting the analysis to just the summer 

8 months. Lin et al. (2004a) assessed the impact of short-term changes in air pollutants on

9 the number of daily neonatal deaths in Sao Paulo, Brazil. The authors observed no 

10 association between ambient levels of O3 and neonatal mortality. 

7.4.10.4 Postneonatal Mortality, 1 Month to 1 Year

11 A number of studies focused on the postneonatal period when examining the effects of O3

12 on infant mortality. Ritz et al. (2006) linked birth and death certificates for infants who 

13 died between 1989 and 2000 to evaluate the influence of outdoor air pollution on infant 

14 death in the South Coast Air Basin of California. The authors examined short- and long15 term exposure periods 2 weeks, 1 month, 2 months, and 6 months before each case 

16 subject’s death and reported no association between ambient levels of O3 and 

17 postneonatal mortality. Woodruff et al. (2008) evaluated the county-level relationship 

18 between cause-specific postneonatal infant mortality and long-term early-life exposure 

19 (first 2 months of life) to air pollutants across the U.S. Similarly, they found no 

20 association between O3 exposure and deaths from respiratory causes. In the U.K., Hajat et 

21 al. (2007) analyzed time-series data of daily infant mortality counts in 10 major cities to 

22 quantify any associations with short-term changes in air pollution. When the results from 

23 the 10 cities were combined there was no relationship between O3 and postneonatal 

24 mortality, even after restricting the analysis to just the summer months. In Ciudad Juarez, 

25 Mexico, Romieu et al. (2004a) examined the daily number of deaths between 1997 and 

26 2001, estimating the modifying effect of SES on the risk of postneonatal mortality. 

27 Ambient O3 concentrations were not related to infant mortality overall, or in any of the 

28 SES groups. In a follow-up study, Carbajal-Arroyo et al. (2011) evaluated the 

29 relationship of 1-h daily max O3 levels with postneonatal infant mortality in the 

30 Mexico City Metropolitan Area between 1997 and 2005. Generally, short-term exposure 

31 to O3 was not significantly related to infant mortality. However, upon estimating the 

32 modifying effect of SES on the risk of postneonatal mortality, the authors found that O3

33 was statistically significantly related to respiratory mortality among those with low SES. 

34 In a separate analysis, the effect of PM10 was evaluated with O3 level quartiles. PM10 

35 alone was related to a significant increase in all-cause mortality. The magnitude of this 
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1 effect remained the same when only the days when O3 was in the lowest quartile were 

2 included in the analyses. However, when only the days when O3 was in the highest 

3 quartile were included in the analyses, the magnitude of the PM10 effect increased 

4 dramatically (OR = 1.06 [95% CI: 0.909, 1.241] for PM10 on days with O3 in lowest 

5 quartile; OR = 1.26 [95% CI: 1.08, 1.47] for PM10 on days with O3 in the highest quartile. 

6 These results suggest that while O3 alone may not have an effect on infant mortality, it 

7 may serve to potentiate the observed effect of PM10 on infant mortality.

8 Tsai et al. (2006) used a case-crossover analysis to examine the relationship between 

9 short-term exposure to air pollution and postneonatal mortality in Kaohsiung, Taiwan 

10 during the period 1994-2000. The risk of postneonatal deaths was 1.023 (95% CI: 0.564,

11 1.858) per 10-ppb increase in 24-h avg O3. The confidence interval for this effect 

12 estimate is very wide, likely due to the small number of infants that died each day, 

13 making it difficult to interpret this result. Several other studies conducted in Asia did not 

14 find any association between O3 concentrations and infant mortality in the postneonatal 

15 period. Ha et al. (2003) conducted a daily time-series study in Seoul, Korea to evaluate 

16 the effect of short-term changes in ambient 8-h O3 concentrations on postneonatal 

17 mortality. Son et al. (2008) examined the relationship between air pollution and 

18 postneonatal mortality from all causes among firstborn infants in Seoul, Korea during 

19 1999-2003. Yang et al. (2006) used a case-crossover analysis to examine the relationship 

20 between air pollution exposure and postneonatal mortality in Taipei, Taiwan for the 

21 period 1994-2000. The authors observed no associations between ambient levels of O3

22 and postneonatal mortality. 

7.4.10.5 Sudden Infant Death Syndrome

23 The strongest evidence for an association between ambient O3 concentrations and SIDS 

24 comes from a study that evaluated the county-level relationship between SIDS and long25 term early-life exposure (first 2 months of life) to air pollutants across the U.S. 

26 (Woodruff et al., 2008). The authors observed a 1.20 (95% CI: 1.09, 1.32) odds ratio for a 

27 10-ppb increase in O3 and deaths from SIDS. There was a monotonic increase in odds of 

28 SIDS for each quartile of O3 exposure compared with the lowest quartile (highest quartile 

29 OR = 1.51; [95% CI: 1.17, 1.96]). In a multi-pollutant model including PM10 or PM2.5, 

30 CO and SO2, the OR for SIDS and O3 was not substantially lower than that found in the 

31 single-pollutant model. When examined by season, the relationship between SIDS deaths 

32 and O3 was generally consistent across seasons with a slight increase for those babies 

33 born in the summer. When stratified by birth weight, the OR for LBW babies was 1.27 

34 (95% CI: 0.95, 1.69) per 10-ppb increase in O3 and the OR for normal weight babies was 

35 1.16 (95% CI: 1.01, 1.32) per 10-ppb increase in O3. 
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1 Conversely, two additional studies reported no association between ambient levels of O3 

2 and SIDS. Ritz et al. (2006) linked birth and death certificates for infants who died 

3 between 1989 and 2000 to evaluate the influence of outdoor air pollution on infant death 

4 in the South Coast Air Basin of California. The authors examined short- and long-term 

5 exposure periods 2 weeks, 1 month, 2 months, and 6 months before each case subject’s 

6 death and reported no association between ambient levels of O3 and SIDS. Dales et al. 

7 (2004) used time-series analyses to compare the daily mortality rates for SIDS and short8 term air pollution concentrations in 12 Canadian cities during the period of 1984-1999. 

9 Increased daily rates of SIDS were associated with previous day increases in the levels of 

10 SO2, NO2, and CO, but not O3 or PM2.5. 

11 Table 7-8 provides a brief overview of the epidemiologic studies of infant mortality. 

12 These studies have focused on short-term exposure windows (e.g., 1-3 days) and long13 term exposure windows (e.g., up to 6 months). Collectively, they provide no evidence for 

Table 7-8 Brief summary of infant mortality studies.

Study Location Mean O3 (ppb)

Exposure 

Assessment Effect Estimatea (95% CI):

Pereira et al. (1998) Sao Paulo, Brazil 1-h max: 

33.8

Citywide avg L0-2: 1.00 (0.99, 1.01)

Diaz et al. (2004) Madrid, Spain 24-h avg: 

11.4

Citywide avg NR

Loomis et al. (1999) Mexico City, 

Mexico

24-h avg: 

44.1

1-h max: 

163.5

1 monitor L0: 0.99 (0.97, 1.02)

L1: 0.99 (0.96, 1.01)

L2: 1.00 (0.98, 1.03)

L3: 1.03 (1.00, 1.05)

L4: 1.01 (0.98, 1.03)

L5: 1.02 (0.99, 1.04)

L0-2: 1.02 (0.99, 1.05)

Ritz et al. (2006) Southern 

California

24-h avg: 21.9-22.1 Nearest Monitor 2 weeks before death: 1.03 (0.93, 1.14)

1 mo before death: NR

2 mo before death: 0.93 (0.89, 0.97)

6 mo before death: NR

Hajat et al. (2007) 10 Cities in the 

UK

24-h avg: 20.5-42.6 Citywide avg L0-2: 1.00 (0.96, 1.06)

Lin et al. (2004a) Sao Paulo, Brazil 24-h avg: 

38.06

Citywide avg L0: 1.00 (0.99, 1.01)

Ha et al. (2003) Seoul, South 

Korea

8-h avg: 

21.2

Citywide avg L0: 0.93 (0.90, 0.96)

Romieu et al. (2004a) Ciudad Juarez, 

Mexico

8-h avg: 

43.43-55.12

Citywide avg L1: 0.96 (0.90, 1.03)

L2: 0.97 (0.91, 1.04)

L0-1 cum: 0.96 (0.89, 1.04)

L0-2 cum: 0.94 (0.87, 1.02)
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Study Location Mean O3 (ppb)

Exposure 

Assessment Effect Estimatea (95% CI):

Carbajal-Arroyo et al. 

(2011)

Mexico City, 

Mexico

1-h max: 

103.0

Citywide avg L0: 1.00 (0.99, 1.00)

L1: 0.99 (0.99, 0.99)

L2: 0.99 (0.99, 1.00)

L0-2: 0.99 (0.99, 1.00)

Son et al. (2008) Seoul, South 

Korea

8-ha avg: 

25.61

Citywide avg L(NR): 0.984 (0.976, 0.992)b

Tsai et al. (2006) Kaohsiung, 

Taiwan

24-h avg: 

23.60

Citywide avg L0-2 cum: 1.02 (0.56, 1.86)

Woodruff et al. (2008) Nationwide, U.S. 24-h avg: 

26.6

County wide avg First 2 mo of life: 1.04 (0.98, 1.10)

Yang et al. (2006) Taipei, Taiwan 24-h avg: 

18.14

Citywide avg L0-2 cum:1.00 (0.62, 1.61)

Dales et al. (2004) 12 Canadian 

cities

24-h: 31.77 Citywide avg L0: NR

L1: NR

L2: NR

L3: NR

L4: NR

L5: NR

Multiday lags of 2-6 days: NR

a

Relative risk of infant mortality per 10 ppb change in O3

b

No increment provided

L0 = Lag 0, L1= Lag 1, L2 = Lag 2, L3 = Lag 3, L4 = Lag 4, L5 = Lag 5, L6 = Lag 6

NR: No quantitative results reported

Table 7-9 Summary of key reproductive and developmental toxicological 

studies.

Study Model

O3

(ppm) Exposure Duration Effects

Sharkhuu et 

al. (2011)

Pregnant mice; 

BALB/c; F;

GD9-18; effects 

in offspring

0.4, 

0.8, or 

1.2

Continuously for 10 

consecutive days

Dams: Decreased number of dams reaching parturition. 

Offspring: (1)-Decreased birth weights. (2)-Decreased rate of 

postnatal growth (body weight). (3)-impaired delayed type 

hypersensitivity.(4)-No effect on humoral immunity. 

(5)-Significantly affected allergic airway inflammation markers 

(eosinophilia, IgE) in female offspring sensitized early in life. 

6-BALF LDH significantly elevated in female offspring.

Bignami et 

al. (1994)

Pregnant CD-1 

dams (PD7-17)

0.4, 0.8 

or 1.2

Continuous Reproductive success was not affected by O3 exposure (PD7-

17, proportion of successful pregnancies, litter size, ex ratio, 

frequency of still birth, or neonatal mortality). Ozone acted as 

a transient anorexigen in pregnant dams.

Haro and

Paz (1993)

Rat dams, 

Exposure over 

the entirety of 

pregnancy;

1.0 12h/day during dark 

cycle

Decreased birth weight and postnatal body weight of offspring 

out to PND 90. Ozone-exposed pups manifested with inverted 

sleep-wake patterns or circadian rhythm phase-shift.

López et al. 

(2008)

Rats; Pregnant 

dams; GD1-

GD18, GD20, 

or GD21.

1.0 (12 h/day, out to either 

GD18, GD20 or GD21)

O3 induced delayed maturation of near term rodent 

bronchioles, with ultra-structural damage to bronchiolar 

epithelium.
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Study Model

O3

(ppm) Exposure Duration Effects

Auten et al. 

(2009)

C57BL/6 

mouse pups

1.0 3 h/day, every other 

day, thrice weekly for 4 

weeks

Postnatal O3 exposure significantly increased lung 

inflammatory cytokine levels; this was further exacerbated 

with gestational PM exposure.

Plopper et al. 

(2007)

Infant rhesus 

monkeys

0.5 Postnatal, PND30-

6month of age, 5 

months of cyclic 

exposure, 5 days O3

followed by 9 days of 

filtered air, 8h/day.

Non-significant increases airway resistance and airway 

responsiveness with O3 or inhaled allergen alone. Allergen + 

O3 produced additive changes in both measures.

Fanucchi et 

al. (2006)

Infant male 

Rhesus 

monkeys, postnatal exposure

0.5 5 months of episodic 

exposure, age 1 monthage 6 months, 5 days 

O3 followed by 9 days 

of filtered air, 8h/day.

Cellular changes and significant structural changes in the 

distal respiratory tract in infant rhesus monkeys exposed to O3

postnatally.

Dell'Omo et 

al. (1995)

CD-1 Mouse 

dams and pups

0.6 6 days before breeding 

to weaning at PND21

Laterality changes in offspring: Ozone exposed pups showed 

a turning preference (left turns) distinct from air exposed 

controls (clockwise turns) as adults.

Santucci et 

al. (2006)

CD-1 Mouse 

dams

0.3 or 

0.6

Dam exposure prior to 

mating through GD17.

Developmental O3 caused increased defensive/submissive 

behavior in offspring. O3 exposed offspring also had significant 

elevations of striatal BDNF and hippocampal NGF v. air 

exposed controls.

Han et al. 

(2011)

Rat; Sprague 

Dawley, M & F; 

PND13

0.6 3 h, BALF examined 

10h after O3 exposure

BALF polymorphonuclear leukocytes and total BALF protein 

were significantly elevated in O3 exposed pups. Lung tissue 

from O3 exposed pups had significant elevations of 

manganese superoxide dismutase (SOD) protein and 

significant decrements of extra-cellular SOD protein.

CamposBedolla et al. 

(2002)

Pregnant Rats; 

Sprague 

Dawley (GD5, 

GD10, or 

GD18)

3.0 1 h on one day of 

gestation, uteri 

collected 16-18 h later

Ozone inhalation modifies the contractile response of the 

pregnant uterus. The O3 exposed pregnant uteri had 

significant increases in the maximum response to acetyl 

choline stimulation at GD5 and 10; they also had a significant 

increase in maximal response to oxytocin at GD 5.

Kavlock et al. 

(1980)

CD-1 mice; 

(pregnancy day 

7-17)

0.4, 0.8 

and 1.2

Continuous, pregnancy 

day 7-17

O3 induced decrements in postnatal body weight gain. When 

O3 was co-administered with sodium salicylate, O3

synergistically increased the rate of pup resorption (1.0 ppm 

GD9-12).

JedlinskaKrakowska 

et al. (2006)

5 month old 

male Wistar 

Hannover rats

3.0 0.5 ppm, 5h/day for 

50 days

Histopathological evidence of impaired spermatogenesis 

(round spermatids/ 21 spermatocytes, giant spermatid cells, 

and focal epithelial desquamation with denudation to the 22 

basement membrane). Vitamin E exposure concomitant with 

O3 protected against pathological changes but Vitamin C did 

not.

7.4.11 Summary and Causal Determination

1 The 2006 O3 AQCD concluded that the limited number of studies that investigated O3

2 demonstrated no associations between O3 and birth outcomes, with the possible exception 

3 of birth defects. The current review included an expanded body of evidence on the 

4 associations between O3 and reproductive and developmental effects. Recent 

5 epidemiologic and toxicological studies provide evidence for an effect of prenatal 

6 exposure to O3 on pulmonary structure and function, including lung function changes in 

7 the newborn, incident asthma, ultrastructural changes in bronchiole development, 
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1 alterations in placental and pup cytokines, and increased pup airway hyper-reactivity. 

2 Also, there is limited toxicological evidence for an effect of prenatal and early life 

3 exposure on central nervous system effects, including laterality, brain morphology, 

4 neurobehavioral abnormalities, and sleep aberration. Recent epidemiologic studies have 

5 begun to explore the effects of O3 on sperm quality, and provide limited evidence for 

6 decrements in sperm concentration, while there is limited toxicological evidence for 

7 testicular degeneration associated with O3. 

8 While the collective evidence for many of the birth outcomes examined is generally 

9 inconsistent (including birth defects), there are several well-designed, well-conducted 

10 studies that indicate an association between O3 and adverse outcomes. For example, as 

11 part of the southern California Children’s Health Study, Salam et al. (2005) observed a 

12 concentration-response relationship of decreasing birth weight with increasing O3

13 concentrations averaged over the entire pregnancy that was clearest above the 30-ppb 

14 level (see Figure 7-4). Similarly, Hansen et al. (2008) utilized fetal ultrasonic 

15 measurements and found a change in ultrasound measurements associated with O3 during 

16 days 31-60 of gestation indicated that increasing O3 concentration decreased an 

17 ultrasound measurement for women living within 2 km of the monitoring site.

18 The weight of evidence does not indicate that prenatal or early life O3 concentrations are 

19 associated with infant mortality. Collectively, there is limited though positive 

20 toxicological evidence for O3-induced developmental effects, including effects on 

21 pulmonary structure and function and central nervous system effects. Limited 

22 epidemiologic evidence for an effect on prenatal O3 exposure on respiratory development 

23 provides coherence with the effects observed in toxicological studies. There is also 

24 limited epidemiologic evidence for an association with O3 concentration and decreased 

25 sperm concentration. A recent toxicological study provides limited evidence for a 

26 possible biological mechanism (histopathology showing impaired spermatogenesis) for 

27 such an association. Additionally, though the evidence for an association between O3

28 concentrations and adverse birth outcomes is generally inconsistent, there are several 

29 influential studies that indicate an association with reduced birth weight and restricted 

30 fetal growth. 

31 Some of the key challenges to interpretation of these study results include the difficulty in 

32 assessing exposure as most studies use existing monitoring networks to estimate 

33 individual exposure to ambient air pollution (see Section 4.6); the inability to control for 

34 potential confounders such as other risk factors that affect birth outcomes (e.g., smoking); 

35 evaluating the exposure window (e.g., trimester) of importance; integrating the results 

36 from both short- and long-term exposure periods; integrating the results across a variety 
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1 of reproductive and developmental outcomes; and limited evidence on the physiological 

2 mechanism of these effects.

3 Taking into consideration the positive evidence for developmental and reproductive 

4 outcomes from toxicological and epidemiological studies, and the few influential birth 

5 outcome studies, the evidence is suggestive of a causal relationship between 

6 exposures to O3 and reproductive and developmental effects.

7.5 Central Nervous System Effects 

7.5.1 Effects on the Brain and Behavior

7 The 2006 O3 AQCD included toxicological evidence that acute exposures to O3 are 

8 associated with alterations in neurotransmitters, motor activity, short and long term 

9 memory, and sleep patterns. Additionally, histological signs of neurodegeneration have 

10 been observed. Reports of headache, dizziness, and irritation of the nose with O3

11 exposure are common complaints in humans, and some behavioral changes in animals 

12 may be related to these symptoms rather than indicative of neurotoxicity. Research in the 

13 area of O3-induced neurotoxicity has notably increased over the past few years, and 

14 recent studies examining the effects of long-term exposure have demonstrated 

15 progressive damage in various regions of the brains of rodents in conjunction with altered 

16 behavior. Evidence from epidemiologic studies has been more limited. A recently 

17 published epidemiologic study examined the association between O3 concentration and 

18 neurobehavioral effects. Chen and Schwartz (2009) utilized data from the NHANES III 

19 cohort to study the relationship between O3 concentrations (mean annual O3

20 concentration 26.5 ppb) and neurobehavioral effects among adults aged 20-59 years. 

21 Annual O3 concentration was determined using inverse distance weighting for county of 

22 residence and adjacent counties (for more information on inverse distance weighting and 

23 other methods for exposure assessment, see Section 4.5.1 and 4.6). The authors observed 

24 an association between annual O3 concentration and tests measuring coding ability 

25 (symbol-digit substitution test) and attention/short-term memory (serial-digit learning 

26 test). Each 10-ppb increase in annual O3 concentration corresponded to an aging-related 

27 cognitive performance decline of 3.5 yr for coding ability and 5.3 years for 

28 attention/short-term memory. These associations persisted in both crude and adjusted 

29 models. There was no association between O3 concentration and reaction time tests. The 

30 authors concluded that overall, there is an association between long-term O3

31 concentration and reduced performance on neurobehavioral tests.
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1 A number of recent toxicological studies demonstrate various perturbations in neurologic 

2 function or histology with long-term exposure to O3, including changes similar to those 

3 observed in neurodegenerative disorders such as Parkinson’s and Alzheimer’s disease 

4 pathologies in relevant regions of the brain (Table 7-10). The central nervous system is 

5 very sensitive to oxidative stress, due in part to its high content of polyunsaturated fatty 

6 acids, high rate of oxygen consumption, and low antioxidant enzyme capacity. Oxidative 

7 stress has been identified as one of the pathophysiological mechanisms underlying 

8 neurodegenerative disease (Simonian and Coyle, 1996), and it is believed to play a role in 

9 altering hippocampal function, which causes cognitive deficits with aging (Vanguilder 

10 and Freeman, 2011). A particularly common finding in studies of O3-exposed rats is lipid 

11 peroxidation in the brain, especially in the hippocampus, which is important for higher 

12 cognitive function including contextual memory acquisition. Performance in passive 

13 avoidance learning tests is impaired when the hippocampus is injured. For example, in a 

14 subchronic study, exposure of rats to 0.25 ppm O3 (4 h/day) for 15-90 days caused a 

15 complex array of responses, including a time-dependent increase in lipid peroxidation 

16 products and immunohistochemical changes in the hippocampus that were correlated 

17 with decrements in passive avoidance behavioral tests (Rivas-Arancibia et al., 2010). 

18 Changes included increased numbers of activated microglia, a sign of inflammation, and 

19 progressive neurodegeneration. Notably, continued exposure tends to bring about 

20 progressive, cumulative damage, as shown by this study (Rivas-Arancibia et al., 2010)

21 and others (Santiago-López et al., 2010; Guevara-Guzmán et al., 2009; Angoa-Pérez et 

22 al., 2006). The effects of O3 on passive avoidance test performance were particularly 

23 evident at 90 days for both short- and long-term memory. The greatest extent of cell loss 

24 was also observed at this time point, whereas lipid peroxidation did not increase much 

25 beyond 60 days of exposure.

26 The substantia nigra is another region of the brain affected by O3, and seems particularly 

27 sensitive to oxidative stress because the metabolism of dopamine, central to its function, 

28 is an oxidative process perturbed by redox imbalance. Oxidative stress has been 

29 implicated in the premature death of substantia nigra dopamine neurons in Parkinson’s 

30 disease. Progressive damage has been found in the substantia nigra of male rats after 15, 

31 30, and 60 days of exposure to 0.25 ppm O3 for 4 h/day. Santiago-López et al. (2010)

32 observed a reduction dopaminergic neurons within the substantia nigra over time, with a 

33 complete loss of normal morphology in the remaining cells and virtually no dopamine 

34 immunoreactivity at 60 days. This was accompanied by an increase in p53 levels and 

35 nuclear translocation, a process associated with programmed cell death. Similarly, 

36 Angoa-Pérez et al. (2006) have shown progressive lipoperoxidation in the substantia 

37 nigra and a decrease in nigral neurons in ovariectomized female rats exposed to 0.25 ppm 

38 O3, 4h/day, for 7 - 60 days. Lipid peroxidation effectively doubled between the 30 and 

39 60 day time points. Total nigral cell number was also diminished to the greatest extent at 
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1 60 days, and cell loss was particularly evident in the tyrosine hydroxylase positive cell 

2 population (90%), indicating a selective loss of dopamine neurons or a loss of dopamine 

3 pathway functionality.

4 The olfactory bulb also undergoes oxidative damage in O3-exposed animals, in some 

5 cases altering olfactory-dependent behavior. Lipid peroxidation was observed in the 

6 olfactory bulbs of ovariectomized female rats exposed to 0.25 ppm O3 (4 h/day) for 30 or 

7 60 days (Guevara-Guzmán et al., 2009). O3 also induced decrements in a selective 

8 olfactory recognition memory test, which were significantly greater at 60 days compared 

9 to 30 days, and the authors note that early deficits in odor perception and memory are 

10 components of human neurodegenerative diseases. The decrements in olfactory memory 

11 did not appear to be due to damaged olfactory perception based on other tests early on, 

12 but by 60 days deficits in olfactory perception had emerged. 

13 Memory deficits and associated morphological changes can be attenuated by 

14 administration of α-tocopherol (Guerrero et al., 1999), taurine (Rivas-Arancibia et al., 

15 2000), and estradiol (Guevara-Guzmán et al., 2009; Angoa-Pérez et al., 2006), all of 

16 which have antioxidant properties. In the study by Angoa-Pérez et al. (2006) described 

17 above, estradiol seemed particularly effective at protecting against lipid peroxidation and 

18 nigral cell loss at 60 days compared to shorter exposure durations. The same was true for 

19 amelioration of decrements in olfactory recognition memory (Guevara-Guzmán et al., 

20 2009), although protection against lipid peroxidation was similar for the 30 and 60 day 

21 exposures.

22 CNS effects have also been demonstrated in adult mice whose only exposure to O3

23 occurred while in utero, a period particularly critical for brain development. Santucci et 

24 al. (2006) investigated behavioral effects and gene expression after in utero exposure of 

25 mice to 0.3 or 0.6 ppm O3. Exposure began 30 days prior to mating and continued 

26 throughout gestation. Testing of adult animals demonstrated increased 

27 defensive/submissive behavior and reduced social investigation were observed in both the 

28 0.3 and 0.6 ppm O3 groups. Changes in gene expression of brain-derived neurotrophic 

29 factor (BDNF, increased in striatum) and nerve growth factor (NGF, decreased in 

30 hippocampus) accompanied these behavioral changes. BDNF and NGF are involved in 

31 neuronal organization and the growth, maintenance, and survival of neurons during early 

32 development and in adulthood. This study and two others using short-term exposures 

33 demonstrate that CNS effects can occur as a result of in utero exposure to O3, and 

34 although the mode of action of these effects is not known, it has been suggested that 

35 circulating lipid peroxidation products may play a role (Boussouar et al., 2009). 

36 Importantly, these CNS effects occurred in rodent models after in utero only exposure to 

37 (semi-) relevant concentrations of O3.
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Table 7-10 Central nervous system effects of long-term ozone exposure 

in rats.

Study Model O3 (ppm) Exposure Duration Effects

Angoa-Pérez et 

al. (2006)

Rat; Wistar; F; 

Weight: 300 g; 

Ovariectomized

0.25 7 to 60 days, 4 h/day, 

5 days/week

Long-term estradiol treatment protected 

against O3-induced oxidative damage to 

nigral dopamine neurons, lipid 

peroxidation, and loss of tyrosine 

hydrolase-immunopositive cells. 

GuevaraGuzmán et al. 

(2009)

Rat; Wistar; F; 

Weight: 264 g; 

Ovariectomized

0.25 30 and 60 days, 4h/day Long-term estradiol treatment protected 

against O3-induced oxidative stress and 

decreases in α and β estrogen receptors 

and dopamine β-hydroxlyase in olfactory 

bulb, and deficits in olfactory social 

recognition memory and chocolate 

recognition.

Rivas-Arancibia 

et al. (2010)

Rat; Wistar; M; 

Weight: 250-300 g

0.25 15 to 90 days, 4h/day Ozone produced significant increases in 

lipid peroxidation in the hippocampus, and 

altered the number of p53 positive 

immunoreactive cells, activated and 

phagocytic microglia, GFAP 

immunoreactive cells, double cortine cells, 

and short- and long-term memoryretention latency

Santiago-López 

et al. (2010)

Rat; Wistar; M; 

Weight: 250-300 g

0.25 15, 30, and 60 days, 

4 h/day

Progressive loss of dopamine reactivity in 

the substantia nigra, along with 

morphological changes. Increased p53 

levels and nuclear translocation.

Santucci et al. 

(2006)

Mice; CD-1; M; 

18 weeks old

0.3; 0.6 Females continuously 

exposed from 30 days 

prior to breeding until 

GD17 

Upon behavioral challenge with another 

male, there was a significant increase in 

defensive and freezing postures and 

decrease in the frequency of nosesniffing. These behavioral changes were 

accompanied by a significant increase in 

BDNF in the striatum and a decrease of 

NGF in the hippocampus.

7.5.2 Summary and Causal Determination 

1 The 2006 O3 AQCD included toxicological evidence that acute exposures to O3 are 

2 associated with alterations in neurotransmitters, motor activity, short and long term 

3 memory, and sleep patterns. Additionally, histological signs of neurodegeneration have 

4 been observed. However, evidence regarding chronic exposure and neurobehavioral 

5 effects was not available. Recent research in the area of O3-induced neurotoxicity has 

6 included several long-term exposure studies. Notably, the first epidemiologic study to 

7 examine the relationship between O3 exposure and neurobehavioral effects observed an 

8 association between annual O3 levels and an aging-related cognitive performance decline 

9 in tests measuring coding ability and attention/short-term memory. This observation is 

10 supported by studies in rodents which demonstrate progressive oxidative stress and 

11 damage in the brain and associated decrements in behavioral tests, including those 

12 measuring memory, after subchronic exposure to 0.25 ppm O3. Additionally, 
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1 neurobehavioral changes are evident in animals whose only exposure to O3 occurred in 

2 utero. Collectively, the limited epidemiologic and toxicological evidence is coherent and 

3 suggestive of a causal relationship between O3 exposure and CNS effects.

7.6 Carcinogenic and Genotoxic Potential of Ozone

7.6.1 Introduction

4 The radiomimetic and clastogenic qualities of O3, combined with its ability to stimulate 

5 proliferation of cells in the respiratory tract, have suggested that O3 could act as a 

6 carcinogen. However, toxicological studies of tumorigenesis in the rodent lung have 

7 yielded mixed and often confusing results, and the epidemiologic evidence is equally 

8 conflicted. The 2006 O3 AQCD concluded that, “the weight of evidence from recent 

9 animal toxicological studies and a very limited number of epidemiologic studies do not 

support ambient O3 as a pulmonary carcinogen”1 10 (U.S. EPA, 2006b). 

11 Multiple epidemiologic studies reported in the 2006 O3 AQCD examined the association 

12 between O3 concentration and cancer. The largest of these studies, by Pope et al. (2002), 

13 included 500,000 adults from the American Cancer Society’s (ACS) Cancer Prevention II 

14 study. In this study, no association was observed between O3 concentration and lung 

15 cancer mortality. The Adventist Health Study of Smog (AHSMOG) also examined the 

16 association between O3 concentration and lung cancer mortality (Abbey et al., 1999). 

17 There was a positive association between O3 concentrations and lung cancer mortality 

18 among men. No association was reported for women. Another study using the AHSMOG 

19 cohort assessed the risk of incident lung cancer (Beeson et al., 1998). Among males, an 

20 association with incidence of lung cancer was observed with increasing O3

21 concentrations. When stratified by smoking status, the association persisted among never 

22 smokers but was null for former smokers. No association was detected for females. The 

23 Six Cities Study examined various air pollutants and mortality but did not specifically 

24 explore the association between O3 concentrations and lung cancer mortality due to low 

25 variability in O3 concentrations across the cities (Dockery et al., 1993). An ecologic study 

26 performed in Sao Paulo City, Brazil examined the correlations between O3 concentrations 

27 in four of the city districts and incident cancer of the larynx and lung reported in 1997 

28 (Pereira et al., 2005). A correlation between the average number of days O3

29 concentrations exceeded air quality standards from 1981 to 1990 and cancer incidence 

30 was present for larynx cancer but not for lung cancer. 

 1 The toxicological evidence is presented in detail in Table 6-18 on page 6-116 of the 1996 O3 AQCD and Table AX5-13 on page 

AX5-43 of the 2006 O3 AQCD.
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Early toxicological research demonstrated lung adenoma1 1 acceleration in mice with daily 

2 exposure to 1 ppm over 15 months (Stokinger, 1962). Later work demonstrated a 

3 significant increase in lung tumor numbers in one strain of mouse (A/J) but not another 

4 after exposure to 0.3-0.8 ppm O3 (Last et al., 1987; Hassett et al., 1985). The A/J mouse 

5 strain is known to have a high incidence of spontaneous adenomas, and further studies 

6 using this strain found a statistically significant increase in lung tumor incidence after a 

7 9-month exposure to 0.5 ppm and incidence and multiplicity after a 5 month exposure to 

8 0.12 ppm with a 4-month recovery period (Witschi et al., 1999). However, these findings 

9 were discounted by the study authors due to the lack of a clear concentration-response, 

10 and results from the Hassett et al. 1985 and Last et al. 1987 studies were retrospectively 

11 deemed spurious based on what appeared to be unusually low spontaneous tumor 

12 incidences in the control groups (Witschi, 1991). A study of carcinogenicity of O3 by the 

13 National Toxicology Program (NTP, 1994) reported increased incidences of 

14 alveolar/bronchiolar adenoma or carcinoma (combined) in female B6C3F1 mice exposed 

15 over 2 years to 1.0 ppm O3, but not 0.12 or .5 ppm. No effect was detected in male mice. 

16 For a lifetime exposure to 0.5 or 1.0 ppm O3, an increase in the number of female mice 

17 with adenomas (but not carcinomas or total neoplasms) was found. The number of total 

18 neoplasms was also unaffected in male mice, but there was a marginally increased 

19 incidence of carcinoma in males exposed to 0.5 and 1.0 ppm. Thus there was equivocal 

20 evidence of carcinogenic activity in male mice and some evidence of carcinogenic 

21 activity of O3 in females. Experimental details of the NTP mouse study are available in 

22 Table 6-19 on page 6-121 (U.S. EPA, 1996o) of the 1996 O3 AQCD.

23 In Fischer-344/N rats (50 of each sex per group), neither a 2-year nor lifetime exposure to 

24 O3 ranging from 0.12 to 1.0 ppm was found to be carcinogenic (Boorman et al., 1994; 

25 NTP, 1994). However, a marginally significant carcinogenic effect of 0.2 ppm O3 was 

26 reported in a study of male Sprague-Dawley rats exposed for 6 months (n = 50) 

27 (Monchaux et al., 1996). These two studies also examined co-carcinogenicity of O3 with 

NNK2 28 (Boorman et al., 1994) or a relatively high dose of radon (Monchaux et al., 1996), 

29 finding no enhancement of NNK related tumors and a slight non-significant increase in 

30 tumor incidence after combined exposure with radon, respectively. Another study 

31 exploring co-carcinogenicity was conducted in hamsters. Not only was there no 

32 enhancement of chemically induced tumors in the peripheral lung or nasal cavity, but 

33 results suggested that O3 could potentially delay or inhibit tumor development (Witschi et 

34 al., 1993). Thus there is no concrete evidence that O3 can act as a co-carcinogen. 

 1 NOTE: Although adenomas are benign, over time they may progress to become malignant, at which point they are called 

adenocarcinomas. Adenocarcinoma is the predominant lung cancer subtype in most countries, and is the only lung cancer found in 

nonsmokers. From page 8-33 of the 1970 O3 AQCD: “No true lung cancers have been reported, however, from experimental 

exposures to either O3 alone or any other combination or ingredient of photochemical oxidants.”

2 4-(N-nitrosomethylamino)-1-(3-pyridyl)-1-butanone
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1 Immune surveillance is an important defense against cancer, and it should be noted that 

2 natural killer (NK) cells, which destroy tumor cells in the lung, appear to be inhibited by 

3 higher concentrations of O3 and either unaffected or stimulated at lower concentrations 

4 (Section 6.2.5.4, Infection and Adaptive Immunity). This aspect of tumorigenesis adds 

5 yet another layer of complexity which may be reflected by conflicting results across 

6 studies.

7 The following sections will examine epidemiologic studies of cancer incidence and 

8 mortality and toxicological studies that have been published since the 2006 O3 AQCD. 

9 An epidemiologic study has been published with cancer as the outcome; most 

10 epidemiologic studies examine markers of exposure. 

7.6.2 Lung Cancer Incidence and Mortality

11 A recent re-analysis of the full ACS CPSII cohort by the Health Effects Institute is the 

12 only epidemiologic study that has explored the association between O3 concentration and 

13 cancer mortality since the last O3 AQCD. Krewski et al. (2009) conducted an extended 

14 follow-up of the cohort (1982-2000). Mean O3 concentration [obtained from the 

15 Aerometric Information Retrieval System (AIRS) for 1980] were 22.91 ppb for the full 

16 year and 30.15 ppb for the summer months (April-September). No association was 

17 reported between lung cancer mortality and O3 concentration (HR = 1.00 [95% CI: 

18 0.96-1.04] per 10 ppb O3). Additionally, no association was observed when the analysis 

19 was restricted to the summer months. There was also no association present in a sub20 analysis of the cohort examining the relationship between O3 concentration and lung 

21 cancer mortality in the Los Angeles area.

22 Since the 2006 O3 AQCD, two toxicological studies have examined potential 

23 carcinogenicity of O3 (Kim and Cho, 2009a, b). Looking across both studies, which used 

24 the same mouse strain as the National Toxicology Program study described above (NTP, 

25 1994), 0.5 ppm O3 alone or in conjunction with chemical tumor inducers did not enhance 

26 lung tumor incidence in males or females. However, a 10% incidence of oviductal 

27 carcinoma was observed in mice exposed to 0.5 ppm O3 for 16 weeks. The implications 

28 of this observation are unclear, particularly in light of the lack of statistical information 

29 reported. Additionally, there is no mention of oviductal carcinoma after 32 weeks of 

30 exposure, and no oviductal carcinoma was observed after one year of exposure. The NTP 

31 study did not report any increase in tumors at extrapulmonary sites.
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7.6.3 DNA Damage

1 The potential for genotoxic effects relating to O3 exposure was predicted from the 

2 radiomimetic properties of O3. The decomposition of O3 in water produces OH and HO2

3 radicals, the same species that are generally considered to be the biologically active 

4 products of ionizing radiation. Ozone has been observed to cause degradation of DNA in 

5 a number of different models and bacterial strains. The toxic effects of O3 have been 

6 generally assumed to be confined to the tissues directly in contact with the gas, such as 

7 the respiratory epithelium. Due to the highly reactive nature of O3, little systemic 

8 absorption is predicted. Zelac et al. (1971a); (1971b), however, reported a significant 

9 increase in chromosome aberrations in peripheral blood lymphocytes from Chinese 

10 hamsters exposed to 0.2 ppm for 5 hours. Other in vivo exposure studies found increased 

11 DNA strand breaks in respiratory cells from guinea pigs (Ferng et al., 1997) and mice 

12 (Bornholdt et al., 2002) but only with exposure to higher concentrations of O3 (1 ppm for 

13 72 hours and 1 or 2 ppm for 90 minutes, respectively). In other studies there were no 

14 observations of chromosomal aberrations in germ cells, but mutagenic effects have been 

15 seen in offspring of mice exposed to 0.2 ppm during gestation (blepharophimosis or 

16 dysplasia of the eyelids). The overall evidence for mutagenic activity from in vitro 

17 studies is positive, and in the National Toxicology Program report described above, O3

18 was found to be mutagenic in Salmonella, with and without S9 metabolic activation. No 

19 recent toxicological studies of DNA damage have become available since the 2006 O3

20 AQCD.

21 A number of epidemiologic studies looked at the association between O3 and DNA and 

22 cellular level damages. These changes may be relevant to mechanisms leading to cancers 

23 development and serve as early indicators of elevated risk of mutagenicity. 

24 Two studies performed in California examined cytogenetic damage in relation to O3

25 exposures. Huen et al. (2006) examined cytogenetic damage among African American 

26 children and their mothers in Oakland, CA. Increased O3 (mean monthly 8-h O3

27 concentrations ranged from about 30 ppb in April to 14 ppb in November) was associated 

28 with increased cytogenetic damage (micronuclei frequency among lymphocytes and 

29 buccal cells) even after adjustment for household/personal smoking status and distance30 weighted traffic density. Chen et al. (2006a) recruited college students at the University 

31 or California, Berkeley who reported never smoking and compared their levels of 

32 cytogenetic damage (micronuclei frequency from buccal cells) in the spring and fall. 

33 Cytogenetic damage was greater in the fall, which the authors attributed to the increase in 

34 O3 over the summer. However, O3 levels over 2, 7, 10, 14, or 30 days (concentrations not 

35 given) before collection of buccal cells did not correlate with cytogenetic damage. 

36 Estimated lifetime O3 exposure was also not correlated with cytogenetic damage. 
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1 Additionally, the authors exposed a subset of the students (n = 15) to 200 ppb O3 for 

2 4 hours while the students exercised intermittently. Ozone was found to be associated 

3 with an increase in cytogenetic damage in degenerated cells but not in normal cells 

4 9-10 days after exposure. Increased cytogenetic damage was also noted in peripheral 

5 blood lymphocytes collected 18 hours after exposure. 

6 A study performed in Mexico recruited 55 male workers working indoors (n = 27) or 

7 outdoors (n = 28) in Mexico City or Puebla, Mexico in order to study the relationship 

8 between O3 and DNA damage (detected from peripheral blood samples using the Comet 

9 assay) (Tovalin et al., 2006). The median estimated daily O3 concentrations were 

10 estimated to be 28.5 ppb for outdoor workers and 5.1 ppb for indoor workers in 

11 Mexico City and 36.1 ppb for outdoor workers and 19.5 ppb for indoor workers in 

12 Puebla. Overall, a positive correlation between O3 levels and DNA damage was 

13 observed. However, when examining the relationship by city and workplace, only DNA 

14 damage in outdoor workers in Mexico City remained correlated with O3 levels. 

15 Three studies examining the relationship between O3 concentration and DNA-level 

16 damage have been performed in Europe. The largest of these studies was the GenAir 

17 case-control study, which was nested within the European Prospective Investigation into 

18 Cancer and Nutrition (EPIC) study, and included individuals recruited between 1993 and 

19 1998 from ten European countries. Only non-smokers (must not have smoked for at least 

20 10 years prior to enrollment) were enrolled in the study. The researchers examined DNA 

21 adduct levels (DNA bonded to cancer-causing chemicals) and their relationship with O3

22 concentrations (concentrations not given) (Peluso et al., 2005). A positive association was 

23 seen between DNA adduct levels and O3 concentrations from 1990-1994 but not O3

24 concentrations from 1995-1999. In adjusted analyses with DNA adduct levels 

25 dichotomized as high and low (detectable versus non-detectable), the OR was 1.97 

26 (95% CI: 1.08, 3.58) when comparing the upper tertile of O3 concentration to the lower 

27 two tertiles. Two other European studies were conducted in Florence, Italy. The most 

28 recent of these enrolled individuals from the EPIC study into a separate study between 

29 March and September of 1999 (Palli et al., 2009). The purpose of the study was to 

30 examine oxidative DNA damage (determined by Comet assay using blood lymphocytes) 

31 in association with varying periods of O3 exposure. The researchers observed that longer 

32 periods of high O3 concentrations (values not given) were more strongly correlated with 

33 oxidative DNA damage than shorter periods of time (i.e., the rho [p-value] was 0.26 

34 [0.03] for 0-10 days and 0.35 [0.002] for 0-90 days). This correlation was stronger among 

35 men compared to women. The correlations for all time periods had p-values <0.05 for ex36 and never-smokers. For current smokers, the correlation was only observed among time 

37 periods ≤ 25 days. When adjusted for age, gender, smoking history, traffic pollution 

38 exposure, period of blood draw, and area of residence, the association between O3
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1 concentrations and oxidative DNA damage was positive for O3 concentrations 0-60 days, 

2 0-75 days, and 0-90 days prior to blood draw. Positive, statistically significant 

3 associations were not observed among shorter time periods. The other study performed in 

4 Florence recruited healthy volunteers who reported being non-smokers or light smokers 

5 (Giovannelli et al., 2006). The estimated O3 concentrations during the study ranged from 

6 approximately 4-40 ppb for 3-day averages, 5-35 ppb for 7-day averages, and 

7 7.5-32.5 ppb for 30-day averages. Ozone concentrations were correlated with DNA 

8 strand breaks (measured from blood lymphocytes) over longer exposure periods (p-value: 

9 0.002 at 30 days, p-value: 0.04 at 7 days; p-value: 0.17 at 3 days). This association was 

10 robust to control for temperature, solar radiation, gender, and age. No association was 

11 seen between O3 concentrations and measures of oxidative DNA damage at 3, 7, or 

12 30 days.

7.6.4 Summary and Causal Determination 

13 The 2006 O3 AQCD reported that evidence did not support ambient O3 as a pulmonary 

14 carcinogen. Since the 2006 O3 AQCD, very few epidemiologic and toxicological studies 

15 have been published that examine O3 as a carcinogen, but collectively, study results 

16 indicate that O3 may contribute to DNA damage. O3 concentrations in most 

17 epidemiologic studies were measured using air monitoring data. For more information on 

18 long-term exposure assessment, see Section 4.6.3.2 Overall, the evidence is inadequate 

19 to determine if a causal relationship exists between ambient O3 exposures and 

20 cancer.

7.7 Mortality

21 A limited number of epidemiologic studies have assessed the relationship between long22 term exposure to O3 and mortality in adults. The 2006 O3 AQCD concluded that an 

23 insufficient amount of evidence existed “to suggest a causal relationship between chronic 

24 O3 exposure and increased risk for mortality in humans” (U.S. EPA, 2006b). In addition 

25 to the infant mortality studies discussed in Section 7.4.10, additional studies have been 

26 conducted among adults since the last review; an ecologic study that finds no association 

27 between mortality and O3, several re-analyses of the ACS cohort, one of which 

28 specifically points to a relationship between long-term O3 exposure and an increased risk 

29 of respiratory mortality, and a study of four cohorts of persons with potentially 

30 predisposing conditions. These studies supplement the evidence from long-term cohort 

31 studies characterized in previous reviews of O3, and are summarized here briefly.
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1 In the Harvard Six Cities Study (Dockery et al., 1993), adjusted mortality rate ratios were

2 examined in relation to long-term mean O3 concentrations in six cities: Topeka, KS; St. 

3 Louis, MO; Portage, WI; Harriman, TN; Steubenville, OH; and Watertown, MA. Mean 

4 O3 concentrations from 1977 to 1985 ranged from 19.7 ppb in Watertown to 28.0 ppb in 

5 Portage. Long-term mean O3 concentrations were not found to be associated with 

6 mortality in the six cities. However, the authors noted that “The small differences in O3

7 levels among the (six) cities limited the power of the study to detect associations between 

8 mortality and O3 levels.” In addition, while total and cardio-pulmonary mortality were 

9 considered in this study, respiratory mortality was not specifically considered.

10 In a subsequent large prospective cohort study of approximately 500,000 U.S. adults, 

11 Pope et al. (2002) examined the effects of long-term exposure to air pollutants on 

12 mortality (American Cancer Society, Cancer Prevention Study II). All-cause, 

13 cardiopulmonary, lung cancer and other mortality risk estimates for long-term O3

14 exposure are shown in Figure 7-5. While consistently positive associations were not 

15 observed between O3 and mortality (effect estimates labeled A in Figure 7-5) , the 

16 mortality risk estimates were larger in magnitude when analyses considered more 

17 accurate exposure metrics, increasing when the entire period was considered (effect 

18 estimates labeled B in Figure 7-5) and becoming marginally significant when the 

19 exposure estimate was restricted to the summer months (July to September; effect 

20 estimates labeled C in Figure 7-5), especially when considering cardiopulmonary deaths. 

21 In contrast, consistent positive and significant effects of PM2.5 were observed for both 

22 lung cancer and cardio-pulmonary mortality.
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Years of Data Collection Number of

Metropolitan Areas

Number of Participants

(in thousands)

1-h max O3 Mean (SD)

A 1980-1981 134 559 47.9 (11.0)

B 1982-1998 119 525 45.5 (7.3)

C 1982-1998 (July – Sept) 134 557 59.7 (12.8)

Source: Reprinted with permission of American Medical Association Pope et al. (2002).

Figure 7-5 Adjusted ozone-mortality relative risk estimates (95% CI) by time 

period of analysis per subject-weighted mean ozone concentration 

in the Cancer Prevention Study II by the American Cancer Society.

1 A study by Abbey et al. (1999) examined the effects of long-term air pollution exposure, 

2 including O3, on all-cause (n = 1,575), cardiopulmonary (n = 1,029), nonmalignant 

3 respiratory (n = 410), and lung cancer (n = 30) mortality in the long-term prospective 

4 Adventist Health Study of Smog (AHSMOG) of 6,338 nonsmoking, non-Hispanic white 

5 individuals living in California. A particular strength of this study was the extensive 

6 effort devoted to assessing long-term air pollution exposures, including interpolation to 

7 residential and work locations from monitoring sites over time and space. No associations 

8 with long-term O3 exposure were observed for all cause, cardiopulmonary, and 

9 nonmalignant respiratory mortality. In a follow-up, Chen et al. (2005) utilized data from 

10 the AHSMOG study and reported no evidence of associations between long-term O3

11 exposure (mean O3 concentration 26.2 ppb) and fatal coronary heart disease. Thus, no 

12 association of chronic O3 exposure with mortality outcomes has been detected in this 

13 study. 

14 Lipfert et al. (2003); (2000) reported positive effects on all-cause mortality for peak O3

15 exposures (95th percentile levels) in the U.S. Veterans Cohort study of approximately 

16 50,000 middle-aged men recruited with a diagnosis of hypertension. The actual analysis 
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1 involved smaller subcohorts based on exposure and mortality follow-up periods. Four 

2 separate exposure periods were associated with three mortality follow-up periods. For 

3 concurrent exposure periods, peak O3 was positively associated with all-cause mortality, 

4 with a 9.4% (95% CI: 0.4, 18.4) excess risk per mean 95th percentile O3 less estimated 

5 background level (not stated). “Peak” refers, in this case, to the 95th percentile of 

6 the hourly measurements, averaged by year and county. In a further analysis, Lipfert et al. 

7 (2003) reported the strongest positive association for concurrent exposure to peak O3 for 

8 the subset of subjects with low diastolic blood pressure during the 1982 to 1988 period. 

9 Two more recent studies of this cohort focused specifically on traffic density (Lipfert et 

10 al., 2006a; 2006b). Lipfert et al. (2006b) concluded that: “Traffic density is seen to be a 

11 significant and robust predictor of survival in this cohort, more so than ambient air 

12 quality, with the possible exception of O3,” reporting a significant O3 effect even with 

13 traffic density included in the model: RR = 1.080 per 40 ppb peak O3 (95% CI: 1.019, 

14 1.146). However, in Lipfert et al. (2006a), which considers only the EPA Speciation 

15 Trends Network (STN) sites, O3 drops to non-significant predictor of total mortality for 

16 this cohort. The authors acknowledge that: “Peak O3 has been important in analyses of 

17 this cohort for previous periods, but in the STN data set, this variable has limited range 

18 and somewhat lower values and its small coefficient of variation results in a relatively 

19 large standard error.” The restriction to subjects near STN sites likely reduced the power 

20 of this analysis, though the size of the remaining subjects considered was not reported in 

21 this paper. In addition, these various Veterans Cohort studies considered only total 

22 mortality, and did not consider mortality on a by-cause basis. 

23 An ecological study in Brisbane, Australia used a geospatial approach to analyze the 

24 association of long-term exposure to gaseous air pollution with cardio-respiratory 

25 mortality, in the period 1996-2004 (Wang et al., 2009c). A generalized estimating 

26 equations model was employed to investigate the impact of NO2, O3 and SO2, but PM 

27 was not addressed. The results indicated that long-term exposure to O3 was not associated 

28 with cardio-respiratory mortality, but the fact that this study considered only one city, and 

29 that the range of O3 exposure across that city (23.7-35.6 ppb) was low and slight in 

30 variation in comparison to the range of other pollutants across the city, limited study 

31 power. In addition, confounding factors (e.g., smoking) could not be addressed at the 

32 individual level in this ecological study. Respiratory mortality was not evaluated 

33 separately.

34 A recent study by Zanobetti and Schwartz examined whether year-to-year variations in 

35 8-h mean daily O3 concentrations for the summer (May-September) around their city36 specific long-term trend were associated with year-to-year variations in mortality around 

37 its long-term trend. This association was examined among Medicare participants with 

38 potentially predisposing conditions, including COPD, diabetes, CHF, and MI, defined as 
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1 patients discharged alive after an emergency admission for one of these four conditions. 

2 The analyses was repeated in 105 cities using available data from 1985 through 2006, and 

3 the results were combined using methods previously employed by these authors 

4 (Zanobetti et al., 2008; Zanobetti and Schwartz, 2007). This study design eliminated 

5 potential confounding by factors that vary across city, which is a common concern in 

6 most air pollution cohort studies, and also avoided both confounding by cross-sectional 

7 factors that vary by city and the short-term factors that confound daily time-series studies, 

8 but are not present in annual analyses. The average 8-h mean daily summer O3

9 concentrations ranged from 15.6 ppb (Honolulu, HI) to 71.4 ppb (Bakersfield, CA) for 

10 the 105 cities. The authors observed associations between yearly fluctuations in summer 

11 O3 concentrations and mortality in each of the four cohorts; the hazard ratios (per 10 ppb 

12 increment) were 1.12 (95% CI: 1.06, 1.17) for the CHF cohort, 1.19 (95% CI 1.12, 1.25) 

13 for the MI cohort, 1.14 (95% CI: 1.10, 1.21) for the diabetes cohort, and 1.14 (95% CI: 

14 1.08, 1.19) for the COPD cohort. A key advantage to this study is that fluctuations from 

15 summer to summer in O3 concentrations around long-term level and trend in a specific 

16 city are unlikely to be correlated with most other predicators of mortality risk; except for 

17 temperature, which was controlled for in the regression. Key limitations of the study were 

18 the inability to control for PM2.5, since it was not reliably measured in these cities until 

19 1999, and the inability to separate specific causes of death (e.g., respiratory, 

20 cardiovascular), since Medicare does not provide the underlying cause of death.

21 In the most recent follow-up analyses of the ACS cohort (Jerrett et al., 2009; Smith et al., 

22 2009a), the effects of long-term exposure to O3 were evaluated alone, as well as in 

23 copollutant models with PM2.5 and components of PM2.5. Jerrett et al. (2009) utilized the 

24 ACS cohort with data from 1977 through 2000 (mean O3 concentration ranged from 33.3 

25 to 104.0 ppb) and subdivided cardiopulmonary deaths into respiratory and cardiovascular, 

26 separately, as opposed to combined into one category, as was done by Pope et al. (2002). 

27 Increases in exposure to O3 were associated with an elevated risk of death from 

28 cardiopulmonary, cardiovascular, ischemic heart disease, and respiratory causes. 

29 Consistent with study hypotheses, inclusion of PM2.5 concentrations measured in 

30 1999-2000 (the earliest years for which it was available) as a copollutant attenuated the 

31 association with O3 for all end points except death from respiratory causes, for which a 

32 significant association persisted (Table 7-11). The association between increased O3

33 concentrations and increased risk of death from respiratory causes was insensitive to the 

34 use of a random-effects survival model allowing for spatial clustering within the 

35 metropolitan area and state of residence, and adjustment for several ecologic variables 

36 considered individually. Subgroup analyses showed that temperature and region of 

37 country, but not sex, age at enrollment, body-mass index, education, or PM2.5

38 concentration, modified the effects of O3 on the risk of death from respiratory causes 

39 (i.e., risks were higher at higher temperature, and in the Southeast, Southwest, and Upper 
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1 Midwest). Ozone threshold analyses indicated that the threshold model was not a better 

2 fit to the data (p >0.05) than a linear representation of the overall O3-mortality 

3 association. Overall, this new analysis indicates that long-term exposure to PM2.5

4 increases risk of cardiac death, while long-term exposure to O3 is specifically associated 

5 with an increased risk of respiratory death, and suggests that combining cardiovascular 

6 and respiratory causes of mortality into one category for analysis may obscure any effect 

7 that O3 may have on respiratory-related causes of mortality.

Table 7-11 Relative risk (and 95% CI) of death attributable to a 10-ppb change 

in the ambient ozone concentration.

Cause of Death O3 (96 MSAs)a O3 (86 MSAs)a O3 +PM2.5 (86 MSAs)a

Any Cause 1.001 (0.996, 1.007) 1.001 (0.996, 1.007) 0.989 (0.981, 0.996)

Cardiopulmonary 1.014 (1.007, 1.022) 1.016 (1.008, 1.024) 0.992 (0.982, 1.003)

Respiratory 1.029 (1.010, 1.048) 1.027 (1.007, 1.046) 1.040 (1.013, 1.067)

Cardiovascular 1.011 (1.003, 1.023) 1.014 (1.005, 1.023) 0.983 (0.971, 0.994)

Ischemic Heart Disease 1.015 (1.003, 1.026) 1.017 (1.006, 1.029) 0.973 (0.958, 0.988)

a

Ozone concentrations were measured from April to September during the years from 1977 to 2000, with follow-up from 1982 to 

2000; changes in the concentration of PM2.5 of 10 μg/m3 were recorded for members of the cohort in 1999 and 2000.

Source: Reprinted with permission of Massachusetts Medical Society (Jerrett et al., 2009).

8 In a similar analysis, Smith et al. (2009a) used data from 66 Metropolitan Statistical 

9 Areas (MSAs) in the ACS cohort to examine the association of O3 concentrations during 

10 the warm season and all-cause and cardiopulmonary mortality. Mortality effects were 

11 estimated in single pollutant and copollutant models, adjusting for two PM2.5 constituents, 

12 sulfate, and EC. When all-cause mortality was investigated, there was a 0.8% (95% CI: 

13 -0.31, 1.9) increase associated with a 10 ppb increase in O3 concentration. This 

14 association was diminished when sulfate or EC were included in the model. There was a 

15 2.48% (95% CI: 0.74, 4.3) increase in cardiopulmonary mortality associated with a 

16 10 ppb increase in O3 concentration. The cardiopulmonary association was robust to 

17 adjustment for sulfate, and diminished, though still positive, after adjustment for EC 

18 (1.63% increase; 95% CI: -0.41, 3.7). Smith et al. (2009a) did not specifically separate 

19 out cardiovascular and respiratory causes of death from the cardiopulmonary category, as 

20 was done by Jerrett et al. (2009).

7.7.1 Summary and Causal Determination

21 The The 2006 O3 AQCD concluded that an insufficient amount of evidence existed “to 

22 suggest a causal relationship between chronic O3 exposure and increased risk for 
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1 mortality in humans” (U.S. EPA, 2006b). Several additional studies have been conducted 

2 since the last review that evaluate cause-specific and total mortality. An ecologic study 

3 conducted in Australia observed no association between cardiopulmonary mortality and 

4 O3 (Wang et al., 2009c). Two reanalyses of the ACS cohort were conducted; one 

5 provides weak evidence for an association with cardiopulmonary mortality (Smith et al., 

6 2009a) while the other specifically points to a relationship between long-term O3

7 exposure and an increased risk of respiratory mortality (Jerrett et al., 2009). Most 

8 recently, a study of four cohorts of Medicare enrollees with potentially predisposing 

9 conditions observed associations between O3 and total mortality among each of the 

10 cohorts (Zanobetti and Schwartz, 2011).

11 When considering the entire body of evidence, there is limited support for an association 

12 with long-term exposure to ambient O3 and total mortality. There is inconsistent evidence 

13 for an association between long-term exposure to ambient O3 and cardiopulmonary 

14 mortality, with several analyses from the ACS cohort reporting some positive 

15 associations (Smith et al., 2009a; Pope et al., 2002) while other studies reported no 

16 association (Wang et al., 2009c; Abbey et al., 1999; Dockery et al., 1993). The strongest 

17 evidence for an association between long-term exposure to ambient O3 concentrations 

18 and mortality is derived from associations reported in the Jerrett et al. (2009) study for 

19 respiratory mortality that remained robust after adjusting for PM2.5 concentrations. 

20 Finally, a recent analysis reported associations of ambient O3 concentrations and total 

21 mortality in potentially at-risk populations in the Medicare Cohort (Zanobetti and 

22 Schwartz, 2011), while earlier studies generally report no associations with total 

23 mortality (Lipfert et al., 2006a; Lipfert et al., 2003; Pope et al., 2002; Abbey et al., 1999; 

24 Dockery et al., 1993). Studies of cardiopulmonary and total mortality provide limited 

25 evidence for an association with long-term exposure to ambient O3 concentrations. The 

26 study by Jerrett et al. (2009) observes an association between long-term exposure to 

27 ambient O3 concentrations and respiratory mortality remained robust after adjusting for 

28 PM2.5 concentrations. Coherence and biological plausibility for this observation is 

29 provided by evidence from epidemiologic, controlled human exposure, and animal 

30 toxicological studies for the effects of short- and long-term exposure to O3 on respiratory 

31 effects (See Sections 6.2 and 7.2). Respiratory mortality is a relatively small portion of 

32 total mortality [about 7.6% of all deaths in 2010 were due to respiratory causes (Murphy 

33 et al., 2012)], thus it is not surprising that the respiratory mortality signal may be difficult 

34 to detect in studies of cardiopulmonary or total mortality. Based on the recent evidence 

35 for respiratory mortality along with limited evidence for total and cardiopulmonary 

36 mortality, the evidence is suggestive of a causal relationship between long-term O3

37 exposures and total mortality.
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7.8 Overall Summary

1 The evidence reviewed in this chapter describes the recent findings regarding the health 

2 effects of long-term exposure to ambient O3 concentrations. Table 7-12 provides an 

3 overview of the causal determinations for each of the health categories evaluated.

Table 7-12 Summary of causal determinations for long-term exposures to 

ozone.

Health Category Causal Determination

Respiratory Effects Likely to be a causal relationship

Cardiovascular Effects Suggestive of a causal relationship

Reproductive and Developmental Effects Suggestive of a causal relationship

Central Nervous System Effects Suggestive of a causal relationship

Carcinogenicity and Genotoxicity Inadequate to infer a causal relationship

Total Mortality Suggestive of a causal relationship
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