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7 INTEGRATED HEALTH EFFECTS OF LONGTERM 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 

6 a duration of approximately 30 days (1 month) or longer. 

7 Conclusions from the 2006 O3 AQCD are summarized briefly at the beginning of each 

8 section, and the evaluation of evidence from recent studies builds upon what was 

9 available during the previous review. For each health outcome (e.g., respiratory disease, 

10 lung function), results are summarized for studies from the specific scientific discipline, 

11 i.e., epidemiologic and toxicological studies. The major sections (i.e. respiratory, 

12 cardiovascular, mortality, reproductive/developmental, cancer) conclude with summaries 

13 of the evidence for the various health outcomes within that category and integration of 

14 the findings that lead to conclusions regarding causality based upon the framework 

15 described in Chapter 1. Determination of causality is made for the overall health effect 

16 category, such as respiratory effects, with coherence and plausibility being based on 

17 evidence from across disciplines and also across the suite of related health outcomes, 

18 including cause-specific mortality. 

7.2 Respiratory Effects 

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

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

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

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

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

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

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

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

27 declines, inflammation, and new asthma development remained inconclusive. 
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1 Several studies (Horak et al., 2002; Frischer et al., 1999) collectively indicated that O3

2 exposure over several summer months was associated with smaller increases in lung 

3 function growth in children. For longer time periods (annual), the definitive analysis in 

4 the Child Health Study (CHS) reported by Gauderman et al. (2004) provided little 

5 evidence that such long-term exposure to ambient O3 was associated with significant 

6 deficits in the growth rate of lung function in children in contrast to the effects observed 

7 with other pollutants such as acid vapor, NO2, and PM2.5. Asthmatic children with 

8 GSTM1 null genotype were found to be more susceptible to the impact of O3 exposure 

9 (over a 12 week study period) on small airways function in Mexico (Romieu et al., 

10 2004a). Limited epidemiologic research examined the relationship between long-term O3

11 exposures and inflammation. Evidence of inflammation and allergic responses consistent 

12 with known effects of O3 exposure (30 day mean) such as increased eosinophil levels 

13 were observed in an Austrian study (Frischer et al., 2001). The cross-sectional surveys 

14 available for the 2006 O3 AQCD detected no associations between long-term O3

15 exposures and asthma prevalence, asthma-related symptoms or allergy to common 

16 aeroallergens in children after controlling for covariates.

17 New evidence presented below reports consistent associations between long-term O3

18 exposure and new-onset asthma related to genotype in U.S. cohorts in multi-community 

19 studies. Related studies report coherent relationships between respiratory symptoms 

20 among asthmatics and long-term O3 exposure. Short-term exposure to O3 is associated 

21 with increases in respiratory symptoms and asthma medication use, as summarized in 

22 Section 6.2.4.2. A new line of evidence reports a positive exposure response relationship 

23 between first asthma hospitalization and long-term O3 exposure. Results from recent 

24 studies examining pulmonary function, inflammation, and allergic responses are also 

25 presented. 

7.2.1 New Onset Asthma

26 Risk for new-onset asthma is related in part to genetic susceptibility, behavioral factors 

27 and environmental exposure (Gilliland et al., 1999). Complex chronic diseases, such as 

28 asthma, are partially the result of a sequence of biochemical reactions involving 

29 exposures to various environmental agents metabolized by a number of different genes 

30 (Conti et al., 2003). Understanding the relation between genetic polymorphisms and 

31 environmental exposure can help identify high-risk subgroups in the population and 

32 provide better insight into pathway mechanisms for these complex diseases. Oxidative 

33 stress likely underlies these mechanistic hypotheses (Gilliland et al., 1999). Susceptibility 

34 genes act through modification of disease risk associated with environmental factors. 

35 Epidemiologic investigation of hypotheses of possible mechanisms involving the gene-
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1 environmental (GxE) interaction involves statistical analysis of these interactions for the 

2 risk of new-onset asthma in children being influenced by exposure to air pollution 

3 (Gauderman, 2002, 2001; Gilliland et al., 1999). 

4 Evidence for the potential importance of genetic susceptibility and behavioral factors on 

5 new onset asthma are provided by several recent studies (Himes et al., 2009; Islam et al., 

6 2008; Li et al., 2008; Hanene et al., 2007; Ercan et al., 2006; Li et al., 2006a; Tamer et 

7 al., 2004; Gilliland et al., 2002). Evidence for a gene-pollution interaction in the 

8 pathogenesis of asthma are supported by recent study findings (Islam et al., 2009; Islam 

9 et al., 2008; Oryszczyn et al., 2007; Lee et al., 2004b; Gilliland et al., 2002). 

10 Evidence for associations between long-term exposure to O3 and new-onset asthma is 

11 provided by new studies from the CHS. Initiated in the early 1990’s, the CHS was 

12 originally designed to examine whether long-term exposure to ambient pollutants was 

13 related to chronic respiratory outcomes in children in 12 communities in southern 

14 California (Peters et al., 1999a; Peters et al., 1999b). About 10 years later, the CHS 

15 inaugurated a series of genetic studies (Gilliland et al., 1999) nested within the CHS 

16 cohort by obtaining biological samples from the study subjects (buccal cells). These new 

17 studies examined the relationship between health outcomes, genetic susceptibility, 

18 behavioral factors and environmental exposure.

19 First, the hypothesis that the functional polymorphisms of HMOX-1 [(GT)n repeat], CAT 

20 (-262C > T -844C > T0, and MNSOD (Ala-9Val) are associated with new-onset asthma 

21 was evaluated, and then whether the effects of these variants varied by exposure to O3

22 (Islam et al., 2008). HMOX1 [heme oxygenase (decycling) 1] is a human gene that 

23 encodes for the enzyme heme oxygenase. Heme oxygenase 1 (HO-1) is an enzyme that 

24 catalyzes the metabolism of heme. The heme iron serves as a source or sink of electrons 

25 during electron transfer or redox chemistry, so the presence of the HMOX1 gene, and 

26 therefore the generation of heme oxygenase, protects against oxidative stress in the body. 

27 The authors observed that functional promoter variants in CAT and HMOX-1 showed 

28 ethnicity-specific associations with new-onset asthma and that oxidant gene protection 

29 was restricted to children living in low-O3 communities.

30 The subjects were drawn from the CHS cohort. Children with a history of asthma or 

31 wheeze were excluded from this analysis. Analyses were restricted to children of 

32 Hispanic (n = 576) or non-Hispanic white ethnicity (n = 1,125). New-onset asthma was 

33 classified as having no prior history of asthma at study entry with subsequent report of 

34 physician-diagnosed asthma at follow-up with the date of onset assigned to be the 

35 midpoint of the interval between the interview date when asthma diagnosis was first 

36 reported and the previous interview date. As a sensitivity analysis, the asthma definition 

37 was restricted to those new-onset asthma cases who also used an inhaler (n = 121). They 
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1 calculated long-term mean pollutant levels (1994 – 2003) to assign exposure to children 

2 in each community for use in the statistical analysis. The effect of ambient air pollution 

3 on the relationship between genetic polymorphism and new-onset asthma was assessed 

4 using models where the community specific average air pollution levels were fitted as a 

5 continuous variable together with the appropriate interaction terms for genes and air 

6 pollutants (Berhane et al., 2004). Cox proportional hazard regression models were fitted 

7 to the data. A stratified analysis for the two independent fourth-grade cohorts of the study 

8 population recruited in 1993 and 1996 was conducted to assess whether the results could 

9 be replicated in independent groups of children. 

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

11 The evidence indicated that the effect of variation in the HMOX-1 gene on risk of new12 onset asthma differed by ambient O3 level. An interaction P value was reported of 0.003 

13 from the hierarchical two stage Cox proportional hazard model fitting the community14 specific O3 and PM10 levels (continuous) and controlling for random effect of the 

15 communities. Average O3 levels showed low correlation with the other monitored 

16 pollutants. The interaction indicated a greater effect (association) of community O3 level 

17 among children with the gene than with children without the gene. Alleles with 23 or 

18 fewer (GT)n repeats are categorized as short (S). The S-allele variant of this protective 

19 enzyme is more readily induced than those with more numerous repeats. The largest 

20 protective effect of the (GT)n repeat polymorphism of HMOX-1 was observed for 

21 children who were S-allele carriers and resided in low-O3 communities with Hazard 

22 Ratio (HR) of 0.44 (95% CI: 0.23, 0.83). The ratio of HR of S-allele carriers who resided 

23 in high O3 communities (HR=0.88; [95% CI: 0.33, 2.34]) was twofold greater than in 

24 those who resided in the low-O3 communities (HR=0.44). The non-parallelism of the two 

25 lines in An interaction p-value of 0.003 was obtained from the hierarchical two stage Cox 

26 proportional hazard model fitting the community specific O3 and controlling for random 

27 effect of the communities. The interaction indicates there is a greater effect (association) 

28 of community O3 level on children with the gene than with children without the gene. 

29 The HRs are off-set as opposed to overlapping in the figure to allow clearer presentation 

30 of the results.

31 Figure 7-1 illustrates the interaction: Children with the S-allele have protection against 

32 the onset of asthma; however, in high- O3 communities, this protection is attenuated. The 

33 results from sensitivity analyses on the two fourth-grade cohorts, and the inhaler 

34 definition for asthma were both consistent with the main results. An analysis related to 

35 children’s participation in sports or time spent outdoors produced the same outcome. No 

36 significant interactions were observed between PM10 or other pollutants and the HMOX -

37 1 gene; quantitative results were not presented. A potential concern for not adjusting for 

38 multiple testing was considered by the authors as not a factor in this analysis because the 
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1 selection of the genes was based on a priori hypotheses defined by a well-studied 

2 biological pathway. Thus in this cohort in southern California, Islam et al. (2008) related 

3 new-onset asthma to O3 exposure in genetically susceptible children.

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

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 effect (association) of 

community O3 level on children with the gene than with children without the gene. The HRs are off-set as opposed to overlapping in 

the figure to allow clearer presentation of the results.

Figure 7-1 Interaction of gene presence and O3 level on the Hazard Ratio (HR) 

of new-onset asthma in the 12 Children’s Health Study 

communities.

4 Related to the findings in Islam et al. (2008) discussed above, Islam et al. (2009)

5 examined whether GSTP1, GSTM1, exercise and O3 exposure have interrelated effects 

6 on the pathogenesis of asthma. A modifying role of air pollution on the association 

7 between Ile105Val and asthma in a cohort of children had been observed (Lee et al., 

8 2004b), but the study did not examine O3 specifically or consider exercise. A primary 

9 conclusion that the authors (Islam et al., 2009) reported was that the common functional 

10 variants of GSTP1 and GSTM1 null genotypes modulate the risk of new onset asthma 

11 during adolescence. Children who had the GSTM1 null genotype were at 1.6-fold (95% 

12 CI: 1.2, 2.2) increased risk of developing new onset asthma compared with those without 

13 the null genotype. Further, the CHS investigators examined the complex interrelationship

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 of antioxidant defenses with asthma risk with increasing doses of O3, resulting from 

2 increasing ventilation associated with vigorous exercise characterized by the number of 

3 team sports played. In an earlier analysis, McConnell et al. (2002) had reported that the 

4 risk of new onset asthma was associated with outdoor exercise, especially in high O3

5 communities but did not consider genetic variants. In this new study, Islam et al. (2009)

6 find a six fold increased risk of asthma (HR=6.15, [95% CI: 2.2, 7.4]) for children who 

7 were homozygous for Ile105, participated in three or more team sports and lived in 

8 high-O3 communities, demonstrating the potential importance of a combination of 

9 genetic variability, O3 exposure and behavior on asthma risk. 

10 Epidemiologic evidence of associations of arginase variants with asthma are limited (Li 

11 et al., 2006a). Asthmatic subjects have higher arginase activity than non-asthmatic 

12 subjects (Morris et al., 2004). NO is a mediator of nitrosative stress synthesized from L13 arginine by nitric oxide synthases. In the CHS, Salam et al. (2009) examined whether 

14 arginase variants (ARG1 and ARG2 genes) were associated with asthma and whether 

15 atopy and exposures to smoking and air pollution influence the associations. The 

16 modifying effect of O3 and atopy on the association between haplotypes and asthma were 

17 evaluated using likelihood ratio tests with appropriate interaction terms. They found that 

18 both ARG1 and ARG2 genetic loci were associated with childhood-onset asthma. The 

19 effect of the ARG1 haplotype varied by the child’s history of atopy and ambient O3. 

20 Among atopic children living in high O3 communities, those carrying the ARG1 

21 haplotype had reduced asthma risk (Odds Ratio [OR] per ARG1h4 haplotype copy = 

22 0.12; [95% CI: 0.04, 0.43]; P heterogeneity across atopy/O3 categories = 0.008). 

23 Further, the CHS presents results examining the relationship of new onset asthma with 

24 traffic-related pollution near homes and schools (McConnell et al., 2010). Asthma risk 

25 increased with modeled traffic-related pollution exposure from roadways near homes and 

26 near schools. The HR was 0.76 (95% CI: 0.38, 1.54) across the range of ambient O3

27 exposure in the communities. With adjustment for school and residential non-freeway 

28 traffic-related exposure, the estimated HR for O3 was 1.01 (95% CI: 0.49, 2.11). Gene 

29 variants were not evaluated in this study.

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

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

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

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

34 evidence.

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

36 (2005) assessed the effects of air pollutants on the risk of asthma. The study population 

37 was recruited from elementary and middle schools within 1 km of air monitoring stations. 
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1 The risk of asthma was related to O3 in the one-pollutant model. The addition of other 

2 pollutants (NOx, CO2, SO2, and PM10), in two-pollutant and three-pollutant models, 

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

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

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

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

7 other pollutants. Both evaluation for asthma symptoms and FEV1 suggested that O3

8 increased asthma prevalence. Clark et al. (2010) investigated the effect of exposure to 

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

10 incidence asthma diagnosis up to 3-4 years of age in a population-based nested case11 control study for all children born in southwestern British Columbia in 1999 and 2000 

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

13 was estimated based on their residential address history using regulatory monitoring data, 

14 land use regression modeling, and proximity to stationary pollutant sources. Daily values 

15 from the three closest monitors within 50 km were used to calculate exposures. Traffic16 related pollutants were associated with the highest risk. Ozone was inversely correlated 

17 with the primary traffic-related pollutants (r = -0.7 to -0.9). The low reliability of asthma 

18 diagnosis in infants makes this study difficult to interpret (Martinez et al., 1995). In a 

19 cross-sectional analysis, Akinbami et al. (2010) examined the association between 

20 chronic exposure to outdoor pollutants (12-month avg levels by county) and asthma 

21 outcomes in a national sample of children ages 3-17 years living in U.S. metropolitan 

22 areas (National Health Interview Survey, N = 34,073). A 5-ppb increase in estimated 8-h 

23 max O3 concentration (annual average) yielded a positive association for both currently 

24 having asthma and for having at least 1 asthma attack in the previous year; while the 

25 adjusted odds ratios for other pollutants were not statistically significant. Models in 

26 which pollutant value ranges were divided into quartiles produced comparable results. 

27 Multi-pollutant models (SO2 and PM) produced similar results. The median value for 

28 12-month avg O3 levels was 39.5 ppb and the IQR was 35.9-43.7 ppb. The adjusted odds 

29 for current asthma for the highest quartile (49.9-59.5 ppb) of estimated O3 exposure was 

30 1.56 (95% CI: 1.15, 2.10) with a positive dose-response relationship apparent from the 

31 lowest quartile to the highest. Thus, this cross-sectional analysis and Hwang et al. (2005)

32 provides further evidence relating O3 exposure and the risk of asthma.

33 The occurrence of bronchitic symptoms among children with asthma was investigated in 

34 the CHS examining the role of gene-environment interactions and long-term O3

35 exposure. Lee et al. (2009b) studied associations of TNF-308 genotype with bronchitis 

36 symptoms among asthmatic children and investigated whether associations vary with 

37 ambient O3 exposure since increased airway TNF may be related to inflammation. 

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

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

2 with TNF-308 GG genotype had a significantly reduced risk of bronchitic symptoms with 

3 low-O3 exposure (OR=0.53; [95% CI: 0.31, 0.91]). The risk was not reduced in children 

4 living in high-O3 communities (OR=1.42; [95% CI: 0.75, 2.70]). The difference in 

5 genotypic effects between low- and high-O3 environments was statistically significant 

6 among asthmatics (P for interaction = 0.01), but insignificant among non-asthmatic 

7 children. Using indicator variables for each category based on genotype and O3 exposure, 

8 Lee et al. (2009b) calculated the effect of TNF-308 GG genotype on the occurrence of 

9 bronchitic symptoms among children with asthma. Figure 7-2 presents adjusted O3 

10 community-specific beta-coefficients plotted against ambient O3 concentration, using 

11 weights proportional to the inverse variance. They further report that they found no 

12 substantial differences in the effect of the GG genotype in asthmatic children in relation 

13 to exposure to PM10, PM2.5, NO2, acid vapor or second-hand smoke exposure. These 

14 results suggest a role of gene-environment interactions such as long-term O3 exposure on

15 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 G-308A genotype on bronchitic 

symptoms among children with asthma in the CHS. Using 

indicator variables for each category based on genotype and O3

exposure, betas were calculated of TNF-308 GG genotype on the 

occurrence of bronchitic symptoms among children with asthma. 
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1 The French Epidemiology study on Genetics and Environment of Asthma (EGEA) 

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

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

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

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

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

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

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

9 lived within 10 km of a monitor, but where 70% of the O3 concentrations were back10 extrapolated values; and (2 [second method]) a validated spatial model that used 

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

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

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

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

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

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

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

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

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

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

21 2.74 (95% CI: 1.68, 4.48) per IQR days of 10-28 (+18) ppb. The effect estimates for SO2

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

23 0.68, 1.29) respectively. Taking into account duration of residence did not change the 

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

25 exposure.

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

27 between asthma and O3, NO2, and PM10. New aspects considered include: 1) 

28 examination of three domains of asthma control (symptoms, exacerbations, and lung 

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

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

31 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 32 for summer (April33 September) O3. The association between asthma control and air pollutants was expressed 

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

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

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

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

38 positively associated with partly controlled and uncontrolled asthma, with a clear gradient 
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1 from controlled, partly controlled (OR=1.53, 95%CI: 1.01, 2.33) and uncontrolled 

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

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

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

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

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

7 1.22, 2.34) and 1.35 (95% CI: 1.13, 1.64) for O3-sum and PM10, respectively). For two 

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

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

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

11 effects of both pollutants are independent. 

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

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

14 defined dichotomously as % predicted FEV1 value < or > =80 (OR=1.35, 95%CI: 0.80, 

15 2.28 for adjusted O3-sum); 2) for symptoms defined as asthma attacks or dyspnoea or 

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

17 95%CI: 1.10, 2.30 for adjusted O3-sum); and for exacerbations defined at least one 

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

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

20 pollutants were more stable and significant when using the integrated measure of asthma 

21 control, this indicates that the results are not driven by one domain. These results support 

22 an effect of long-term exposure to O3 on asthma control in adulthood in subjects with 

23 pre-existing asthma.

24 The interrelationships between variants in catalase (CAT) and myeloperoxidase (MPO) 

25 genes, ambient pollutants, and acute respiratory illness were investigated in a national 

26 U.S. cohort (Wenten et al., 2009). Health information, air pollution, and incident 

27 respiratory-related school absences were ascertained in January-June 1996 for 1,136 

28 Hispanic and non-Hispanic white U.S. elementary schoolchildren as part of the 

29 prospective Air Pollution and Absence Study, a population based cohort study conducted 

30 as part of the CHS. A related earlier study (Gilliland et al., 2001), which was discussed in 

31 the 2006 O3 AQCD, examined the effects of ambient air pollution on school absenteeism 

32 due to respiratory illness without a genetic aspect to the study. In a new study Wenten et 

33 al. (2009) hypothesized that variation in the level or function of these enzymes would 

34 modulate respiratory illness risk, especially under high levels of oxidative stress. The 

35 joint effect of these two genes on respiratory illness was examined. Risk of respiratory36 related school absences was elevated for children with the CAT (G/G) and MPO (G/A or 

37 A/A) genes (relative risk = 1.35, [95% CI: 1.03, 1.77]; P-interaction = 0.005). To assess 
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1 effects of long-term average levels of O3 on acute effects, communities were divided into 

2 high and low exposure groups by median levels (46.9 ppb O3). The epistatic effect of 

3 CAT and MPO variants was evident in communities exhibiting high ambient O3 levels 

4 (P-interaction = 0.03). The association of respiratory-illness absences with functional 

5 variants in CAT and MPO that differ by air pollution levels illustrates the need to 

6 consider genetic epistasis in assessing gene-environment interactions. In high O3

7 communities, CAT/MPO genotypes that resulted in decreased oxidative stress were 

8 associated with a decreased risk of respiratory related school absences compared with the 

9 CAT/MPO wild-type genotype (Relative Risk [RR] = 0.42, [95% CI: 0.20, 0.89]).

7.2.2 Asthma Hospital Admissions and ED Visits

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

25 466.1) hospitalization and two measures of long-term pollutant exposure (the month prior 

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

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

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

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

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

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

32 California Health Interview Survey data from subjects ages 1 to 65+ who reported 

33 physician-diagnosed asthma (n = 1,502). Subjects were assigned annual average 

34 concentrations for O3 based on residential ZIP code and the closet air monitoring station 
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1 within 8 km but did not have data on duration of residence. Multi-pollutant models for O3

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

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

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

5 2.11] per 10 ppb) for all ages. Positive associations were obtained for symptoms but 95% 

6 CIs included null values. Associations for symptoms for adults (ages 18 +) were observed 

7 (OR=1.40; [95% CI: 1.02, 1.91] per 10 ppb).

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

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

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

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

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

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

14 ED visit or hospitalization because of asthma during the past 12 months. This cross15 sectional study reports an OR of 3.34 (95% CI: 1.01, 11.09) for poorly controlled asthma 

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

17 level to those below that level. Multi-pollutant (PM) analysis produced similar results.

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

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

20 2008b). This study investigated the association between chronic exposure to O3 and 

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

22 following a birth cohort of 1,204,396 eligible births born in New York State during 1995-

23 1999 to first asthma admission or until 31 December 2000. There were 10,429 (0.87%) 

24 children admitted to the hospital for asthma between 1 and 6 years of age. The asthma 

25 hospitalization rate in New York State in 1993 was 2.87 per 1,000 (Lin et al., 1999). 

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

27 chronic O3 exposure: (1) mean concentration during the follow-up period (41.06 ppb); (2) 

28 mean concentration during the O3 season (50.62 ppb); and (3) proportion of follow-up 

29 days with O3 levels >70 ppb. In this study the authors aimed to predict the risk of having 

30 asthma admissions in a birth cohort, but the time to the first admission in children that is 

31 usually analyzed in survival models was not their primary interest. The effects of 

32 co-pollutants were assessed and controlled for using the Air Quality Index (AQI). 

33 Interaction terms were used to assess potential effect modifications. A positive 

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

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

36 to develop asthma severe enough to be admitted to the hospital. The various factors were 

37 examined and differences were found for younger children (1-2 years), poor 

38 neighborhoods, Medicaid/self-paid births, geographic region and others. As shown in 
Draft – Do Not Cite or Quote 7-13 September 2011

1 Adjusted for child’s sex, age, birth weight, and gestational age; maternal race, ethnicity, age, education, 

2 insurance, and smoking status during pregnancy; and regional poverty level and temperature. ORs by low, 

3 medium, and high exposure are shown for New York City (NYC: low [37.3 ppb], 

4 medium [37.3 - 38.11] ppb, high [38.11 + ppb]) and other New York State regions (Other 

5 NYS regions: low [42.58 ppb], medium [42.58-45.06 ppb], high [45.06+ ppb]) for first 

6 asthma hospital admission.

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

8 admissions were significantly associated with increased O3 levels for all chronic 

9 exposure indicators (ORs, 1.16-1.68). When estimating the O3 effect using the 

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

11 hospital admissions with an IQR (2.51%) increase in O3. A proportional hazards model 

12 for the New York City data was run as a sensitivity analysis and it yielded similar results 

13 between asthma admissions and chronic exposure to O3 (Cox model: HR = 1.14, [95% 

14 CI: 1.124, 1.155] is similar to logistic model results: OR = 1.16 (95% CI: 1.15, 1.17) 

15 (Lin, 2010). Thus, this study provides evidence associating long-term O3 exposure to first 

16 asthma hospital admission in a concentration-response relationship. 

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.

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

concentration during the entire follow-up period.
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7.2.3 Pulmonary Structure and Function

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 four- to 

9 eightfold 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 

11 influence the association between ambient air pollutant exposures and lung function 

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

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

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

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

16 associated with differences in susceptibility of children for lung function growth deficits 

17 associated with NO2, PM10, PM2.5, elemental carbon, organic carbon, and O3. The 

18 negative effects of air pollutants were largely observed within participants who had a 

19 particular GSS haplotype. The effects ranged from -124.2 to -149.1 mL for FEV1, -92.9 

20 to -126.7 mL for FVC and -193.9 to -277.9 mL/s for MMEF for all pollutants except O3, 

21 for which some positive associations were reported: 25.9 mL for FEV1; 0.1 mL for FVC, 

22 and 166.5 mL/s for MMEF. Ozone was associated with larger decreases in lung function 

23 in children without this haplotype, when compared to the other pollutants with values of -

24 76.6 mL for FEV1, -17.2 mL for FVC, and -200.3 mL/s for MMEF, but only the 

25 association with MMEF was statistically significant.

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

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

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

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

30 Subjects with a large ratio were less likely to have decreases in FEF75 and FEF25-75 for a 

31 given estimated lifetime exposure to O3. Kinney and Lippmann (2000) examined 72 

32 nonsmoking adults (mean age 20 years) from the second-year class of students at the U.S. 

33 Military Academy in West Point, NY, and reported results that appear to be consistent 

34 with a decline in lung function that may in part be due to O3 exposures over a period of 

35 several summer months. Ilhorst et al. (2004) examined 2,153 children with a median age 

36 of 7.6 years and reported pulmonary function results which indicated that significantly 

37 lower FVC and FEV1 increases were associated with higher O3 exposures over the 
Draft – Do Not Cite or Quote 7-15 September 2011

1 medium-term of several summer months, but not over several months in the winter. 

2 Semi-annual mean O3 concentrations ranged from 22 to 54 ppb during the summer 

3 months and 4 to 36 ppb during the winter months. However, over the longer-term 

4 3.5-year period Ilhorst et al. (2004) found no associations between increases in lung 

5 function and mean summer months O3 levels for FVC and FEV1, in contrast to the 

6 significant medium-term effects. Frischer et al. (1999) showed results similar to the 

7 Ilhorst et al. (2004) study. 

8 Mortimer et al. (2008a, b) examined the association of prenatal and lifetime exposures to 

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

33 just markers for lifetime or current exposures.

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

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

36 inhomogeneity and eNO were measured in 241 unsedated, sleeping neonates (age = 

37 5 weeks). Consistent with the previous studies, no association was found for prenatal 

38 exposure to O3 and lung function.
Draft – Do Not Cite or Quote 7-16 September 2011

1 In a cross-sectional study of adults, Qian et al. (2005) examined the association of long2 term exposure to O3 and PM10 with pulmonary function from data of 10,240 middle-aged 

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

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

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

6 with measures of pulmonary function. 

7 To determine the extent to which long-term exposure to outdoor air pollution accelerates 

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

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

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

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

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

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

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

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

16 Mexico City (n = 3,170) who were 8 years of age at the beginning of the study, Rojas17 Martinez et al. (2007) evaluated the association between long-term exposure to O3, PM10

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

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

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

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

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

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

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

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

26 from copollutant models were not substantially different than single pollutant models, 

27 independent effects for pollutants could not be estimated accurately because the traffic28 related pollutants were correlated. To reduce exposure misclassification, 

29 microenvironmental and personal exposure assessments were conducted in a randomly 

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

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

32 monitoring stations. 

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

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

35 exposures over several summer months on lung function in children, i.e., reduced lung 

36 function growth being associated with higher ambient O3 levels. Longer-term studies, 

37 investigating the association of chronic O3 exposure on lung function such as the CHS, 
Draft – Do Not Cite or Quote 7-17 September 2011

1 were inconclusive. Short-term O3 exposure studies presented in Section 6.2.1.2 provide a 

2 cumulative body of epidemiologic evidence that strongly supports associations between 

3 ambient O3 exposure and decrements in lung function among children. For new studies 

4 of long-term O3 exposure relationship to pulmonary function, one study, where O3 and 

5 other pollutant levels were higher (90 ppb at high end of the range) than those in the 

6 CHS, observes a relationship between O3 concentration and pulmonary function declines 

7 in school-aged children. Two studies of adult cohorts provide mixed results where long8 term exposures were at the high end of the range with levels of 49.5 ppb in one study and 

9 27 ppb IQR in the other. Thus there is little new evidence to build upon the very limited 

10 studies from the 2006 O3 AQCD.

7.2.3.1 Pulmonary Structure and Function: Evidence from 

Toxicological Studies

11 As reviewed in the 1996 and 2006 O3 AQCDs, there are both qualitative and quantitative 

12 uncertainties in the extrapolation of data generated by rodent toxicology studies to the 

13 understanding of health effects observed in humans, as documented by epidemiologic and 

14 controlled exposure studies. Chief among these data extrapolation issues are the 

15 differences between rodent and human respiratory physiology, cellular makeup, 

16 dosimetry, and morphometry (see Chapter 5). However, important evidence is available 

17 from O3-inhalation studies performed in nonhuman primates whose respiratory system 

18 most closely resembles that of the human. A long series of studies have used nonhuman 

19 primates to examine the effect of O3 alone or in combination with an inhaled allergen, 

20 house dust mite antigen, on morphology and lung function. These studies, by Plopper and 

21 colleagues, have demonstrated changes in pulmonary function and airway morphology in 

22 adult and infant nonhuman primates repeatedly exposed to environmentally relevant 

23 concentrations of O3 (Joad et al., 2008; Carey et al., 2007; Plopper et al., 2007; Fanucchi 

24 et al., 2006; Joad et al., 2006; Evans et al., 2004; Larson et al., 2004; Tran et al., 2004; 

25 Evans et al., 2003b; Schelegle et al., 2003; Fanucchi et al., 2000; Hyde et al., 1989; 

26 Harkema et al., 1987a; Harkema et al., 1987b; Fujinaka et al., 1985). The findings of 

27 these nonhuman primate studies have also been observed in rodent studies discussed at 

28 the end of this section and included in Table 7-1.

29 Since publication of the 1996 and 2006 O3 AQCDs, the initial observations in adult 

30 nonhuman primates have been expanded in a series of experiments using infant rhesus 

31 monkeys repeatedly exposed to 0.5 ppm O3 starting at 1 month of age (Plopper et al., 

32 2007). Many of the observations found in adult monkeys have also been noted in infant 

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

34 infant monkeys has not been reported. In terms of pulmonary function changes, after 
Draft – Do Not Cite or Quote 7-18 September 2011

1 several episodic exposures of infant monkeys to O3 (each cycle: 5 days of 0.5 ppm O3 at 

2 8 h/day, followed by 9 days of filtered air exposure), they observed more than a doubling 

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

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

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

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

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

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

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

10 of an O3-induced change in airway responsiveness supports the biologic plausibility of 

11 long-term exposure to O3 contributing to the effects of asthma in children. To understand 

12 which conducting airways and inflammatory mechanisms are involved in O3-induced 

13 airway hyperresponsiveness in the infant rhesus monkey, a follow-up study examined 

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

15 to morphometrically evaluate the response of bronchi and respiratory bronchioles to 

16 methacholine, (a bronchoconstricting agent commonly used to evaluate airway 

17 responsiveness in asthmatics), the investigators observed differential effects for the two 

18 airway sizes. While episodic exposure to O3 alone (0.5 ppm) had little effect on ex vivo 

19 airway responsiveness in bronchi and respiratory bronchioles, exposure to dust mite 

20 antigen alone produced airway hyperresponsiveness in the large bronchi, whereas O3 + 

21 antigen produced significant increases in airway hyperresponsiveness only in the 

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

23 responsiveness occurs predominantly in the smaller bronchioles, where dosimetric 

24 models indicate the dose would be higher.

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

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

27 morphological changes, including a significant fourfold increase in eosinophils, (a cell 

28 type important in allergic asthma), in the bronchoalveolar lavage of infant monkeys 

29 exposed to O3 alone. Thus, these studies demonstrate both functional and cellular 

30 changes in the lung of infant monkeys after cyclic exposure to 0.5 ppm O3. This 

31 concentration, while higher than those used in controlled human exposure studies, 

32 provides relevant information to understanding the adverse effects of ambient O3

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

34 are available from these nonhuman primate studies.

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

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

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

38 in the infant rhesus monkey. Exposure to O3 alone (5 days of 0.5 ppm O3 at 8 h/day, 
Draft – Do Not Cite or Quote 7-19 September 2011

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

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

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

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

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

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

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

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

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

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

11 basement membrane zone morphometry. These latter effects are significant because of 

12 their potential contribution to airway obstruction and airway hyperresponsiveness which 

13 are central features of asthma.

14 As noted above, a significant increase in airway responsiveness to inhaled histamine 

15 occurred in infant rhesus monkeys exposed to O3 + house dust mite antigen, but not to O3

16 alone (Schelegle et al., 2003). To study the underlying mechanisms of this airway 

17 hyperresponsiveness, these investigators evaluated the effect of exposure to O3 alone and 

18 in combination with (+) antigen on non-specific airway responsiveness to methacholine 

19 at different airway generations. After exposure to filtered air, O3, antigen, or O3 + 

20 antigen, the bronchi and respiratory bronchioles of 6-month-old monkeys were 

21 challenged ex vivo with methacholine. Exposure to O3 alone had no significant effect on 

22 airway responsiveness to methacholine in either airway, whereas O3 + antigen produced 

23 a significant increase in airway responsiveness in the respiratory bronchioles but not the 

24 larger bronchi.

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

26 asthma, the effect of exposure to O3 alone or O3 + antigen on immune system parameters 

27 was also examined in infant rhesus monkeys. Mast cells, which contribute to asthma via 

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

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

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

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

32 allergic asthma (Van Winkle et al., 2010). Increases in CD4+ and CD8+ lymphocytes 

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

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

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

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

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

38 CD25+ cells, it did alter the anatomic distribution of CD25+ cells within the airways. 
Draft – Do Not Cite or Quote 7-20 September 2011

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

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

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

4 epithelial cells are thought to play a contributing role to airway hyperresponsiveness. As 

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

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

7 al., 2004). Whereas, a significant reduction in airway innervation occurred after exposure 

8 to O3 alone, a significantly greater reduction occurred in monkeys exposed to O3 + 

9 antigen. This reduction in overall airway innervation was accompanied, however, by an 

10 increase in the abundance of protein gene product 9.5, a nonspecific neural marker. 

11 Significant increases in protein gene product 9.5 were still observed in O3 alone- and O3

12 + antigen-exposed infant monkeys after a 6-month recovery protocol (Kajekar et al., 

13 2007). Thus, in addition to structural, immune, and inflammatory effects, exposure to O3

14 produces alterations in airway innervation which may contribute to O3-induced 

15 exacerbation of asthma. 

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

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

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

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

20 cross links that appear to be irreversible (Reiser et al., 1987). Generally changes in 

21 collagen content have been observed in rats exposed to 0.5 ppm O3 or higher, although 

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

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

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

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

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

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

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

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

30 evaluated males. It is unclear what the long-term effects of these structural changes may 

31 be. A number of studies indicate that structural changes in the respiratory system are 

32 persistent or irreversible. For example, O3-induced hyperplasia was still evident in the 

33 nasal epithelia of rats 13 weeks after recovery from 0.5 ppm O3 exposure (Harkema et 

34 al., 1999). In a study of episodic exposure to 0.25 ppm O3, Chang et al. (1992) observed 

35 no reversal of basement membrane thickening in rat lungs up to 17 weeks post-exposure. 

36 Episodic exposure (0.25 ppm O3, every other month) of monkeys induced equivalent 

37 morphological changes compared to continuously exposed animals, even though they 
Draft – Do Not Cite or Quote 7-21 September 2011

1 were exposed for half the time and evaluation occurred a month after exposure ceased as 

2 opposed to immediately (Tyler et al., 1988). 
Draft – Do Not Cite or Quote 7-22 September 2011

Table 7-1 Respiratory effects in nonhuman primates and rodents resulting 

from long-term O3 exposure

Study Model O3 (ppm) Exposure Duration Effects

Catalano et al. (1995a; 

1995b); Chang et al. 

(1995); Harkema et al. 

(1997a; 1997b; 1994); 

Last et al. (1994); 

Pinkerton et al. (1995); 

Plopper et al.(1994); 

Stockstill et al. (1995);

Pinkerton et al. (1998)

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 in rats exposed to 0.12 ppm O3. Other 

effects (e.g., mucous cell metaplasia in the nose and 

mild fibrotic response in the parenchyma, increased 

collagen in CAR of females) observed at 0.5 to 1.0 

ppm. Some morphometric changes such as 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 concentration-response 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 

mo for 18 mos

Episodic group 

evaluated 1 mo post 

exposure

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 week sold

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.
Draft – Do Not Cite or Quote 7-23 September 2011

Study Model O3 (ppm) Exposure Duration Effects

Schelegle et al. (2003); Monkey; Rhesus,

30 days old*

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.

Larson et al. (2004) Monkey;Macaca 

mulatta, 30 days 

old*

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 old*

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.

* 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 

15 studies discussed earlier in Section 6.2.3.2 show consistent associations of O3 exposure 

16 and increased airway inflammation and oxidative stress. Further discussion of the 

17 mechanisms underlying inflammation and oxidative stress responses can be found in 

18 Section 5.3.3. Though the majority of recent studies focus on short-term exposures, 
Draft – Do Not Cite or Quote 7-24 September 2011

1 several epidemiologic and toxicology studies of long-term exposure add to observations 

2 of O3-induced inflammation and injury.

3 Inflammatory markers and peak expiratory pulmonary function were examined in 37 

4 allergic children with physician-diagnosed mild persistent asthma in a highly polluted 

5 urban area in Italy and then again 7 days after relocation to a rural location with 

6 significantly lower pollutant levels (Renzetti et al., 2009). The authors observed a 

7 fourfold decrease in nasal eosinophils and a statistically significant decrease in fractional 

8 exhaled nitric oxide along with an improvement in lower airway function. Several 

9 pollutants were examined, including PM10, NO2, and O3, though pollutant-specific 

10 results were not presented. These results are consistent with studies showing that traffic11 related exposures are associated with increased airway inflammation and reduced lung 

12 function in children with asthma and contribute to the notion that this negative influence 

13 may be rapidly reversible. Exhaled NO (eNO) has been shown to be a useful biomarker 

14 for airway inflammation in large population-based studies (Linn et al., 2009). Thus, while 

15 the time scale of 7 days between examinations for eNO needs to be evaluated for 

16 appropriateness, the results suggest that inflammatory responses are reduced when O3

17 levels are decreased.

18 Chest radiographs (CXR) of 249 children in Mexico City who were chronically exposed 

19 to O3 and PM2.5 were analyzed by Calderón-Garcidueñas et al. (2006). They reported an 

20 association between chronic exposures to O3 and other pollutants and a significant 

21 increase in abnormal CXR’s and lung CTs suggestive of a bronchiolar, peribronchiolar, 

22 and/or alveolar duct inflammatory process, in clinically healthy children with no risk 

23 factors for lung disease. These CXR and CT results should be viewed with caution 

24 because it is difficult to attribute effects to O3 exposure.

25 In a cross-sectional study, Wood et al. (2009) examined the association of outdoor air 

26 pollution with respiratory phenotype (PiZZ type) in alpha 1-Antitrypsin deficiency (α27 ATD) from the U.K. α-ATD registry. In total, 304 PiZZ subjects underwent full lung 

28 function testing and quantitative high-resolution computed tomography to identify the 

29 presence and severity of COPD – emphysema. Mean annual air pollution data for 2006 

30 was matched to the location of patients’ houses and used in regression models to identify 

31 phenotypic associations with pollution controlling for covariates. Relative trends in O3

32 levels were assessed to validate use of a single year’s data to indicate long-term exposure 

33 and validation; data showed good correlations between modeled and measured data 

34 (Stedman and Kent, 2008). Regression models showed that estimated higher exposure to 

35 O3 exposure was associated with worse gas transfer and more severe emphysema, albeit 

36 accounting for only a small proportion of the lung function variability. This suggests that 

37 a gene-specific group demonstrates a long-term O3 exposure effect.
Draft – Do Not Cite or Quote 7-25 September 2011

1 The similarities of nonhuman primates to humans make them attractive models in which 

2 to study the effects of O3 on the respiratory tract. The nasal mucous membranes, which 

3 protect the more distal regions of the respiratory tract, are susceptible to injury from O3. 

4 Carey et al. (2007) conducted a study of O3 exposure in infant rhesus macaques, whose 

5 nasal airways closely resemble that of humans. Monkeys were exposed either acutely for 

6 5 days (8 h/day) to 0.5 ppm O3, or episodically for several biweekly cycles alternating 

7 5 days of 0.5 ppm O3 with 9 days of filtered air (0 ppm O3), designed to mimic human 

8 exposure (70 days total). All monkeys acutely exposed to O3 had moderate to marked 

9 necrotizing rhinitis, with focal regions of epitheliar exfoliation, numerous infiltrating 

10 neutrophils, and some eosinophils. The distribution, character, and severity of lesions in 

11 episodically exposed monkeys were similar to that of acutely exposed animals. Neither 

12 group exhibited mucous cell metaplasia proximal to the lesions, a protective adaptation 

13 observed in adult monkeys exposed continuously to 0.3 ppm O3 in another study 

14 (Harkema et al., 1987a). Adult monkeys also exhibit attenuation of inflammatory 

15 responses with continued daily exposure (Harkema et al., 1987a), but inflammation did 

16 not resolve over time in young episodically exposed monkeys(Carey et al., 2011). 

17 Inflammation in conducting airways has also been observed in rats chronically exposed to 

18 O3. Using an agar-based technique to fill the alveoli so that only the rat bronchi are 

19 lavaged, a 90-day exposure of rats to 0.8 ppm O3 (8 h/day) elicited significantly elevated 

20 pro-inflammatory eicosanoids PGE2 and 12-HETE in the conducting airway compared to 

21 filtered air-exposed rats (Schmelzer et al., 2006).

7.2.5 Allergic Responses

22 The association of air pollutants with childhood respiratory allergies was examined in the 

23 U.S. using the 1999-2005 National Health Interview Survey of approximately 70,000 

24 children, and ambient air pollution data from the U.S. EPA, with monitors within 20 

25 miles of each child’s residential block (Parker et al., 2009). The authors examined the 

26 associations between the reporting of respiratory allergy or hay fever and medium-term 

27 exposure to O3 over several summer months, controlling for demographic and geographic 

28 factors. Increased respiratory allergy/hay fever was associated with increased O3 levels 

29 (adjusted OR per 10 ppb = 1.20; [95% CI: 1.15, 1.26]). These associations persisted after 

30 stratification by urban-rural status, inclusion of multiple pollutants (O3, SO2, NO2, PM), 

31 and definition of exposure by differing exposure radii; smaller samples within 5 miles of 

32 monitors were remarkably similar to the primary results. No associations between the 

33 other pollutants and the reporting of respiratory allergy/hay fever were apparent. 

34 Ramadour et al. (2000) reported no relationship between O3 levels and rhinitis symptoms 

35 and hay fever. Hwang et al. (2006) report the prevalence of allergic rhinitis (adjusted OR 
Draft – Do Not Cite or Quote 7-26 September 2011

1 per 10 ppb = 1.05; [95% CI: 0.98, 1.12]) in a large cross-sectional study in Taiwan. In a 

2 large cross-sectional study in France, Penard-Morand et al. (2005) reported a positive 

3 relationship between lifetime allergic rhinitis and O3 exposure in a two-pollutant model 

4 with NO2. These studies related positive outcomes of allergic response and O3 exposure 

5 but with variable strength for the effect estimates. A toxicological study reported that 

6 five weeks of continuous exposure to 0.4 ppm O3 (but not 0.1 or 0.2 ppm O3) augmented 

7 sneezing and nasal secretions in a guinea pig model of nasal allergy (Iijima and 

8 Kobayashi, 2004). Nasal eosinophils, which participate in allergic disease and 

9 inflammation, and allergic antibody levels in serum were also elevated by exposure to 

10 concentrations as low as 0.2 ppm (Iijima and Kobayashi, 2004).

11 Nasal eosinophils were observed to decrease by fourfold in 37 atopic, mildly asthmatic 

12 children 7 days after relocation from a highly polluted urban area in Italy to a rural 

13 location with significantly lower pollutant levels (Renzetti et al., 2009). Inflammatory 

14 and allergic effects of O3 exposure (30 day mean) such as increased eosinophil levels 

15 were observed in children in an Austrian study (Frischer et al., 2001). Episodic exposure 

16 of infant rhesus monkeys to 0.5 ppm O3 for 5 months appears to significantly increase the 

17 number and proportion of eosinophils in the blood and airways (lavage) [protocol 

18 described above in 7.2.3.1 for Fanucchi et al. (2006)] (Maniar-Hew et al., 2011). These 

19 changes were not evident at 1 year of age (6 months after O3 exposure ceased). Increased 

20 eosinophils levels have also been observed after acute or prolonged exposures to O3 in 

21 adult bonnet and rhesus monkeys (Hyde et al., 1992; Eustis et al., 1981). 

22 Total IgE levels were related to air pollution levels in 369 adult asthmatics in five French 

23 centers using generalized estimated equations (GEE) as part of the EGEA study described 

24 earlier (Rage et al., 2009b). Geostatistical models were performed on 4×4 km grids to 

25 assess individual outdoor air pollution exposure that was assigned to subject’s home 

26 address. Ozone concentrations were positively related to total IgE levels and an increase 

27 of 5 ppb of O3 resulted in an increase of 20.4% (95% CI: 3.0, 40.7) in total IgE levels. 

28 Nearly 75% of the subjects were atopic. In two-pollutant models including O3 and NO2, 

29 the O3 effect estimate was decreased by 25% while the NO2 effect estimate was decreased 

30 by 57%. Associations were not sensitive to adjustment for covariates or the season of IgE 

31 measurements. These cross-sectional results suggest that exposure to O3 may increase 

32 total IgE in adult asthmatics. 

33 Although very few toxicological studies of long-term exposure examining allergy are 

34 available, short-term exposure studies in rodents and nonhuman primates demonstrate 

35 allergic skewing of immune responses and enhanced IgE production. Due to the 

36 persistent nature of these responses, the short-term toxicological evidence lends 
Draft – Do Not Cite or Quote 7-27 September 2011

1 biological plausibility to the limited epidemiologic findings of an association between 

2 long-term O3 exposure and allergic outcomes.

7.2.6 Host Defense

3 Short-term exposures to O3 have been shown to cause decreases in host defenses against 

4 infectious lung disease in animal models. However, acute O3-induced suppression of 

5 alveolar phagocytosis and immune functions observed in animals appears to be transient 

6 and attenuated with continuous or repeated exposures. Chronic exposures (weeks, 

7 months) of 0.1 ppm do not cause greater effects on infectivity than short exposures, due 

8 to defense parameters becoming reestablished with prolonged exposures, although 

9 chronic exposure has been shown to slow alveolar clearance. No detrimental effects were 

10 seen with a 120-day exposure to 0.5 ppm O3 on acute lung injury from influenza virus 

11 administered immediately before O3 exposure started. However, O3 was shown to 

12 increase the severity of postinfluenzal alveolitis and lung parenchymal changes (Jakab 

13 and Bassett, 1990). Little new evidence has become available to address the effects of 

14 long-term exposure on host defense mechanisms. However, a recent study by Maniar15 Hew et al. (2011) demonstrated that the immune system of infant rhesus monkeys 

episodically exposed to 0.5 ppm O3 for 5 months1 16 appeared to be altered in ways that 

17 could diminish host defenses. Reduced numbers of circulating leukocytes were observed, 

18 particularly polymorphonuclear leukocytes (PMNs) and lymphocytes, which were 

19 decreased in the blood and airways (bronchoalveolar lavage). These changes did not 

20 persist at 1 year of age (6 months postexposure); rather, increased numbers of monocytes 

21 were observed at that time point. Challenge with LPS, a bacterial ligand that activates 

22 monocytes and other innate immune cells, elicited lower responses in O3-exposed 

23 animals even though the relevant reactive cell population was increased. This was 

24 observed in both an in vivo inhalation challenge and an ex vivo challenge of peripheral 

25 blood mononuclear cells. Thus a decreased ability to respond to pathogenic signals was 

26 observed six months after O3 exposure ceased, in both the lungs and periphery.

7.2.7 Respiratory Mortality

27 A limited number of epidemiologic studies have assessed the relationship between long28 term exposure to O3 and mortality. The 2006 O3 AQCD concluded that an insufficient 

29 amount of evidence existed “to suggest a causal relationship between chronic O3

30 exposure and increased risk for mortality in humans” (U.S. EPA, 2006b). Though total 

 1 Exposure protocol is described above in Section 7.2.3.1 for Fanucchi et al. (2006)
Draft – Do Not Cite or Quote 7-28 September 2011

1 and cardio-pulmonary mortality were considered in these studies, respiratory mortality 

2 was not specifically considered. In the most recent follow-up analysis of the ACS cohort 

3 (Jerrett et al., 2009), cardiopulmonary deaths were subdivided into respiratory and 

4 cardiovascular, separately, as opposed to combined in the Pope et al. (2002) work. A 10-

5 ppb increment in exposure to O3 elevated the risk of death from respiratory causes and 

6 this effect was robust to the inclusion of PM2.5. The association between increased O3

7 concentrations and increased risk of death from respiratory causes was insensitive to the 

8 use of a random-effects survival model allowing for spatial clustering within the 

9 metropolitan area and state of residence, and to adjustment for several ecologic variables 

10 considered individually. Additionally, a recent study (Zanobetti and Schwartz, In Press)

11 observed an association between long-term exposure to O3 and elevated risk of mortality 

12 among Medicare enrollees that had previously experienced an emergency hospital 

13 admission due to COPD.

7.2.8 Summary and Causal Determination

14 The epidemiologic studies reviewed in the 2006 O3 AQCD detected no associations 

15 between long-term (annual) O3 exposures and asthma-related symptoms, asthma 

16 prevalence, or allergy to common aeroallergens among children after controlling for 

17 covariates. Little evidence was available to relate long-term exposure to current ambient 

18 O3 concentrations to deficits in the growth rate of lung function in children. Additionally, 

19 limited evidence was available evaluating the relationship between long-term O3 levels 

20 and pulmonary inflammation and other endpoints. From toxicological studies, it appeared 

21 that O3-induced inflammation tapered off during long-term exposures, but that 

22 hyperplastic and fibrotic changes remained elevated and in some cases even worsened 

23 after a postexposure period in clean air. Episodic exposures were also known to cause 

24 more severe pulmonary morphologic changes than continuous exposure (U.S. EPA, 

25 2006b).

26 The new epidemiologic evidence base consists of studies using a variety of designs and 

27 analysis methods evaluating the relationship between long-term annual measures of 

28 exposure to ambient O3 and measures of respiratory morbidity conducted by different 

29 research groups in different locations. See Table 7-2 for O3 concentrations associated 

30 with selected studies. The positive results from various designs and locations support an 

31 association between long-term O3 concentrations and respiratory morbidity. 

32 New studies examined the relationship between long-term O3 exposure and new onset 

33 asthma in children. Studies have provided evidence for a relationship between different 

34 genetic variants (HMOX, GST, ARG) that, in combination with O3 exposure, are related 
Draft – Do Not Cite or Quote 7-29 September 2011

1 to new onset asthma (Islam et al., 2009; Salam et al., 2009; Islam et al., 2008). These 

2 studies involve two separate cohorts in 12 California communities of the CHS. These 

3 prospective cohort studies represent strong evidence because they are methodologically 

4 rigorous epidemiology studies. The stratified analysis for the two independent fourth5 grade cohorts of the study population recruited in 1993 and 1996 yielded consistent 

6 results and provided replication in independent groups of children. 

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

Akinbami et al. (2010); current asthma

United States 12 month median 39.8 IQR 35.9 to 43.7

Hwang et al. (2005); prevalence of asthma

Taiwan 

Mean 23.14 Range 18.65 to 

31.17

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. 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

Salam et al. (2009); childhood onset asthma; CHS O3 greater than or less than 50 ppb 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 See left

Moore et al. (2008); asthma 

hospital admissions; South Coast Basin Median 87.8 ppb Range 28.6 to 199.9

Meng et al. (2010); 

asthma ED visits or hospitalizations; 

San Joaquin Valley, CA

Median 30.3 ppb 

25-75% range 

27.1 to 34.0

Lee et al. (2009b); 

bronchitic symptoms in asthmatic children; CHS Above and below 50 ppb See left

Rage et al. (2009b); 

asthma severity; five French cities Mean 30 ppb 25th-75th 21-36 

Jacquemin et al. (In Press); 

asthma control in adults; five French cities Median 46.9 ppb; 25th-75th41-52

Wenten et al. (2009); 

respiratory school absence, U.S. Median 46.9 ppb; 10:00 a.m. – 6:00 p.m. Min-Max 27.6-65.3

7 Studies using a cross-sectional design provide support for a relationship between long8 term O3 exposure and health effects in asthmatics. A long-term O3 exposure study relates 

9 bronchitic symptoms to TNF-308 genotype asthmatic children with ambient O3 exposure 

10 in the CHS (Lee et al., 2009b). A study in five French cities reports effects on asthma 

11 severity related to long-term O3 exposure (Rage et al., 2009a). A follow-up study of this 

12 cohort (Jacquemin et al., In Press) supports an effect of long-term O3-sum exposure on 

13 asthma control in adulthood in subjects with pre-existing asthma. Akinbami et al. (2010)

14 and Hwang et al. (2005) provides further evidence relating O3 exposures and the risk of 

15 asthma. For the respiratory health of a cohort based on the general U.S. population, risk 
Draft – Do Not Cite or Quote 7-30 September 2011

1 of respiratory-related school absences was elevated for children with the CAT and MPO 

2 variant genes related to communities with high ambient O3 levels (Wenten et al., 2009).

3 Chronic O3 exposure was related to first childhood asthma hospital admissions in a 

4 positive concentration-response relationship in a New York State birth cohort (Lin et al., 

5 2008b). A separate hospitalization cross-sectional study in San Joaquin Valley in 

6 California reports similar findings (Meng et al., 2010). Another study relates asthma 

7 hospital admissions to quarterly average O3 in the South Coast Air Basin of California 

8 (Moore et al., 2008).

9 Information from toxicological studies indicates that long term exposure to O3 during 

10 gestation or development can result in irreversible morphological changes in the lung, 

11 which in turn can influence pulmonary function. Studies by Plopper and colleagues have 

12 demonstrated changes in pulmonary function and airway morphology in adult and infant 

13 nonhuman primates repeatedly exposed to environmentally relevant concentrations of O3

14 (Fanucchi et al., 2006; Joad et al., 2006; Schelegle et al., 2003; Harkema et al., 1987b). 

15 This nonhuman primate evidence of an O3-induced change in airway responsiveness 

16 supports the biologic plausibility of long term exposure to O3 contributing to the adverse 

17 effects of asthma in children. Results from epidemiologic studies examining long-term 

18 O3 exposure and pulmonary function effects are inconclusive with some new studies 

19 relating effects at higher exposure levels. The results from the CHS described in the 2006 

20 O3 AQCD remain the definitive line of evidence. Other cross-sectional studies provide 

21 mixed results. 
Draft – Do Not Cite or Quote 7-31 September 2011

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

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)

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

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)

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

Rage et al. (2009a)

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

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

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)

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)

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)

The highest quartile is shown for all results.

NA = not available

1 Several studies (see Table 7-3) provide results from studies that adjusted for potential 

2 confounders, presenting results for both O3 and PM (single and multipollutant models) as 

3 well as other pollutants where PM effects were not provided. As shown in the table, O3

4 associations are generally robust to adjustment for potential confounding by PM.

5 The 2006 O3 AQCD concluded that the extensive human clinical and animal 

6 toxicological evidence, together with the limited epidemiologic evidence available, 
Draft – Do Not Cite or Quote 7-32 September 2011

1 suggests a causal role for short-term O3 exposure in inflammatory responses in the 

2 airways. Though the majority of recent studies focus on short-term exposures, several 

3 epidemiologic and toxicological studies of long-term exposure add to observations of O3-

4 induced inflammation and injury. Toxicological studies in rodents and nonhuman 

5 primates indicate that chronic O3 exposure causes structural changes in the respiratory 

6 tract, and simulated seasonal exposure studies suggest that such exposures might have 

7 cumulative impacts. The strongest epidemiologic evidence for a relationship between 

8 long-term O3 exposure and respiratory morbidity is provided by new studies that 

9 demonstrate associations between long-term measures of O3 exposure and new-onset 

10 asthma in children and increased respiratory symptom effects in asthmatics. While there 

11 are currently a limited number of studies in this data base, the U.S. multi-community 

12 prospective cohort studies are methodologically rigorous epidemiologic studies. Asthma 

13 risk is related to complex relationships between genetic variability, environmental O3

14 exposure, and behavior. The genes, evaluated in these studies, are both key candidates in 

15 the oxidative stress pathway and have been shown to play an important role in asthma 

16 development. Reduced risk for asthma development is reported in some studies in 

17 children living in low- O3 communities. Mean O3 concentrations in the studies (10:00 

18 a.m. to 6:00 p.m.) ranged from 28.6 to 45.5 ppb in low O3 communities

19 (mean = 38.4 ppb) and from 46.5 to 64.9 ppb in high O3 communities (mean = 55.2 ppb). 

20 These CHS multi-community studies form a foundation for the evidence base in which 

21 findings for several genes indicate the breath of the evidence across different gene 

22 variants. The several other studies with different designs, analysis, locations and 

23 researchers provide a cumulative collective body of evidence informing these 

24 relationships. The other studies in the new data base provide coherent evidence for long25 term O3 exposure and respiratory morbidity effects such as first asthma hospitalization 

26 and respiratory symptoms in asthmatics. Studies considering other pollutants provide data 

27 suggesting that the effects related to O3 are independent from potential effects of the 

28 other pollutants. Some studies provide evidence for a positive concentration-response 

29 relationship. Short-term studies provide supportive evidence with increases in respiratory 

30 symptoms and asthma medication use, hospital admissions and ED visits for all 

31 respiratory outcomes and asthma, and decrements in lung function in children. The above 

32 discussion of the recent epidemiologic and toxicological data base provides a compelling 

33 case to support the hypothesis that a relationship exists between long-term exposure to 

34 ambient O3 and measures of respiratory morbidity. The 2006 O3 AQCD concluded the 

35 evidence was suggestive but inconclusive at that time. The new epidemiological data 

36 base, combined with toxicological studies in rodents and nonhuman primates, 

37 provides biologically plausible evidence that there is likely to be causal 

38 relationship between long-term exposure to O3 and respiratory morbidity.
Draft – Do Not Cite or Quote 7-33 September 2011

7.3 Cardiovascular Effects

7.3.1 Cardiovascular Disease

7.3.1.1 Cardiovascular Epidemiology

1 Long-term exposure to O3 and its effects on cardiovascular morbidity were not 

2 considered in the 2006 O3 AQCD. However, recent studies have assessed the chronic 

3 effects of O3 exposure on cardiovascular morbidity (Chuang et al., 2011; Forbes et al., 

4 2009a; Chen et al., 2007). The association between O3 exposure and markers of lipid 

5 peroxidation and antioxidant capacity was examined among 120 nonsmoking healthy 

6 college students, aged 18-22 years, from the University of California, Berkeley (Feb-Jun 

7 2002) (Chen et al., 2007). By design, students were chosen from geographic areas so they 

8 had experienced different levels of O3 over their lifetimes and during recent summer 

9 vacation in either greater Los Angeles (LA) or the San Francisco Bay Area (SF). A 

10 marker of lipid peroxidation, 8-isoprostane (8-iso-PGF) in plasma, was assessed. This 

11 marker is formed continuously under normal physiological conditions but has been found 

12 at elevated concentrations in response to environmental exposures. A marker of overall 

13 antioxidant capacity, ferric reducing ability of plasma (FRAP), was also measured. The 

14 lifetime O3 exposure estimates (estimated monthly average) did not show much overlap 

15 between the two geographic areas [median (range): LA, 42.9 ppb (28.5-65.3); SF, 26.9 

16 ppb (17.6-33.5)]. Estimated lifetime O3 exposure was related to 8-iso-PGF [β = 0.025 

17 (pg/mL)/8-h ppb O3, p = 0.0007]. For the 17-ppb cumulative lifetime O3 exposure 

18 difference between LA and SF participants, there was a 17.41-pg/mL (95% CI: 15.43, 

19 19.39) increase in 8-iso-PGF. No evidence of association was observed between lifetime 

20 O3 exposure and FRAP [β = -2.21 (pg/mL)/8-h ppb O3, p = 0.45]. The authors note that 

21 O3 was highly correlated with PM10-2.5 and NO2 in this study population; however, their 

22 inclusion in the O3 models did not substantially modify the magnitude of the associations 

23 with O3. Because the lifetime exposure results were supported by shorter-term exposure 

24 results from analyses considering O3 concentrations up to 30 days prior to sampling, the 

25 authors conclude that persistent exposure to O3 can lead to sustained oxidative stress and 

26 increased lipid peroxidation. However, because there was not much overlap in lifetime 

27 O3 exposure estimates between LA and SF, it is possible that the risk estimates involving 

28 the lifetime O3 exposures could be confounded by unmeasured factors related to other 

29 differences between the two cities. 

30 Forbes et al. (2009a) used the annual average exposures to assess the relationship 

31 between chronic ambient air pollution and levels of fibrinogen and C-reactive protein 
Draft – Do Not Cite or Quote 7-34 September 2011

1 (CRP) in a cross-sectional study conducted in England. Data were collected from the 

2 Health Survey of England for 1994, 1998, and 2003. The sampling strategy was designed 

3 to obtain a representative sample of the English population; however, due to small group 

4 sizes, only data from white ethnic groups were analyzed. For analyses, the annual 

5 concentrations of O3 were averaged for the year of data collection and the previous year 

6 with the exception of 1994 (because pollutant data were not available for 1993). Median 

7 O3 concentrations were 26.7 ppb, 25.4 ppb, and 28 ppb for 1994, 1998, and 2003, 

8 respectively. Year specific adjusted effect estimates were created and combined in a 

9 meta-analysis. No evidence of association was observed for O3 and levels of fibrinogen 

10 or CRP (e.g., the combined estimates for the percent change in fibrinogen and CRP for a 

11 10 ppb increase in O3 were -0.28 [95% CI: -2.43, 1.92] and -3.05 [95% CI: -16.10, 

12 12.02], respectively). 

13 A study was performed in Taiwan to examine the association between long-term O3

14 concentrations and blood pressure and blood markers using the Social Environment and 

15 Biomarkers of Aging Study (SEBAS) (Chuang et al., 2011). Individuals included in the 

16 study were 54 years of age and older. The mean annual O3 concentration during the study 

17 period was 22.95 ppb (SD 6.76 ppb). Positive associations were observed between O3

18 concentrations and both systolic and diastolic blood pressure [changes in systolic and 

19 diastolic blood pressure were 21.51mmHg (95% CI: 16.90, 26.13) and 20.56 mmHg 

20 (95% CI: 18.14, 22.97) per 8.95 ppb increase in O3, respectively). Increased O3

21 concentrations were also associated with increased levels of total cholesterol, fasting 

22 glucose, hemoglobin A1c, and neutrophils. No associations were observed between O3

23 concentrations and triglyceride and IL-6 levels. The observed associations were reduced 

24 when other pollutants were added to the models. Further research will be important for 

25 understanding the effects, if any, of chronic O3 exposure on cardiovascular morbidity 

26 risk.

7.3.1.2 Cardiovascular Toxicology

27 Three new studies have investigated the cardiovascular effects of long-term exposure to 

28 O3 in animal models (See Table 7-4 for study details). In addition to the short-term 

29 exposure effects described in Section 6.3.3, a recent study found that O3 exposure in 

30 genetically hyperlipidemic mice enhanced aortic atherosclerotic lesion area compared to 

31 air exposed controls (Chuang et al., 2009). Chuang et al. (2009) not only provided 

32 evidence for increased atherogenesis in susceptible mice, but also reported an elevated 

33 vascular inflammatory and redox state in wild-type mice and infant primates 

34 (Section 6.3.3.2). This study is compelling in that it identifies biochemical and cellular 

35 events responsible for transducing the airway epithelial reactions of O3 into 
Draft – Do Not Cite or Quote 7-35 September 2011

1 proinflammatory responses that are apparent in the extrapulmonary vasculature (Cole and 

2 Freeman, 2009). 

3 Another recent study provides further evidence for increased vascular inflammation and 

4 oxidation and long term effects in the extrapulmonary space. Rats episodically exposed to 

5 O3 for 16 weeks presented marked increases in gene expression of biomarkers of 

6 oxidative stress, thrombosis, vasoconstriction, and proteolysis (Kodavanti et al., 2011). 

7 Ozone exposure upregulated aortic mRNA expression of heme oxygenase-1 (HO-1), 

8 tissue plasminogen activator (tPA), plasminogen activator inhibitor-1 (PAI-1), von 

9 Willebrand factor (vWf), thrombomodulin, endothelial nitric oxide synthase (eNOS), 

10 endothelin-1 (ET-1), matrix metalloprotease-2 (MMP-2), matrix metalloprotease-3 

11 (MMP-3), and tissue inhibitor of matrix metalloprotease-2 (TIMP-2). In addition, O3

12 exposure depleted some cardiac mitochondrial phospholipid fatty acids (C16:0 and 

13 C18:1), which may be the result of oxidative modifications. The authors speculate that 

14 oxidatively modified lipids and proteins produced in the lung and heart promote vascular 

15 pathology through activation of lectin-like oxidized-low density lipoprotein receptor-1 

16 (LOX-1). Activated LOX-1 induces expression of a number of the biomarkers induced by 

17 O3 exposure and is considered pro-atherogenic. Both LOX-1 mRNA and protein were 

18 increased in mouse aorta after O3 exposure. This study provides a possible pathway and 

19 further support to the observed O3 induced atherosclerosis.

20 Vascular occlusion resulting from atherosclerosis can block blood flow through vessels 

21 causing ischemia. The restoration of blood flow or reperfusion can cause injury to the 

22 tissue from subsequent inflammation and oxidative damage. Ozone exposure enhanced 

23 the sensitivity to myocardial ischemia-reperfusion (I/R) injury in rats while increasing 

24 oxidative stress levels and pro-inflammatory mediators and decreasing production of anti25 inflammatory proteins (Perepu et al., 2010). Both long- and short-term O3 exposure 

26 decreased the left ventricular developed pressure, rate of change of pressure 

27 development, and rate of change of pressure decay and increased left ventricular end 

28 diastolic pressure in isolated perfused hearts (Section 6.3.3.2 for short-term exposure 

29 discussion). In this ex vivo heart model, O3 induced oxidative stress by decreasing SOD 

30 enzyme activity and increasing malondialdehyde levels. Ozone also elicited a 

31 proinflammatory state evident by an increase in TNF-α and a decrease in the anti32 inflammatory cytokine IL-10. The authors conclude that O3 exposure will result in a 

33 greater I/R injury.
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Table 7-4 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 antiinflammatory proteins.

No previous studies investigated cardiovascular effects from long-term exposure to O3.

7.3.2 Cardiac 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, In Press)

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. A study on O3 and 

21 cardiovascular mortality reported no association after adjustment for PM2.5 levels. 
Draft – Do Not Cite or Quote 7-37 September 2011

1 Further epidemiologic studies on cardiovascular morbidity and mortality after long-term 

2 exposure have not been published. 

3 Toxicological evidence on long-term O3 exposure is also limited but three strong 

4 toxicological studies have been published since the previous AQCD. These studies 

5 provide evidence for O3 enhanced atherosclerosis and I/R injury, corresponding with 

6 development of a systemic oxidative, proinflammatory environment. Further discussion 

7 of the mechanisms that may lead to cardiovascular effects can be found in Section 5.3.8. 

8 Although questions exist for how O3 inhalation causes systemic effects, a recent study 

9 proposes a mechanism for development of vascular pathology that involves activation of 

10 LOX-1 by O3 oxidized lipids and proteins. This activation may also be responsible for O3

11 induced changes in genes involved in proteolysis, thrombosis, and vasoconstriction. 

12 Taking into consideration the findings of toxicological studies, and the emerging 

13 evidence from epidemiologic studies, the generally limited body of evidence is 

14 suggestive of a causal relationship between long-term exposures to O3 and 

15 cardiovascular effects. 

7.4 Reproductive and Developmental Effects 

16 Although the body of literature is growing, the research focusing on adverse birth 

17 outcomes is small. Among these studies, various measures of birth weight and fetal 

18 growth, such as low birth weight (LBW), small for gestational age (SGA), and 

19 intrauterine growth restriction (IUGR), and preterm birth (<37-week gestation; [PTB]) 

20 have received more attention in air pollution research, while congenital malformations 

21 are less studied. There are also new studies on reproductive and developmental effects.

22 Infants and fetal development processes may be particularly susceptible to O3-induced 

23 health effects, and although the physical mechanisms are not fully understood, several 

24 hypotheses have been proposed; these include: oxidative stress, systemic inflammation, 

25 vascular dysfunction and impaired immune function (Section 5.3). Study of these 

26 outcomes can be difficult given the need for detailed exposure data and potential 

27 residential movement of mothers during pregnancy. Air pollution epidemiologic studies 

28 reviewed in the 2006 O3 AQCD examined impacts on birth-related endpoints, including 

29 intrauterine, perinatal, postneonatal, and infant deaths; premature births; intrauterine 

30 growth retardation; very low birth weight (weight <1,500 grams) and low birth weight 

31 (weight <2,500 grams); and birth defects. However, in the limited number of studies that 

32 investigated O3, no associations were found between O3 and birth outcomes, with the 

33 possible exception of birth defects. 
Draft – Do Not Cite or Quote 7-38 September 2011

1 Several recent articles have reviewed methodological issues relating to the study of 

2 outdoor air pollution and adverse birth outcomes (Chen et al., 2010a; Woodruff et al., 

3 2009; Ritz and Wilhelm, 2008; Slama et al., 2008). Some of the key challenges to 

4 interpretation of these study results include the difficulty in assessing exposure as most 

5 studies use existing monitoring networks to estimate individual exposure to ambient air 

6 pollution; the inability to control for potential confounders such as other risk factors that 

7 affect birth outcomes (e.g., smoking); evaluating the exposure window (e.g., trimester) of 

8 importance; and limited evidence on the physiological mechanism of these effects (Ritz 

9 and Wilhelm, 2008; Slama et al., 2008). Recently, an international collaboration was 

10 formed to better understand the relationships between air pollution and adverse birth 

11 outcomes and to examine some of these methodological issues through standardized 

12 parallel analyses in datasets from different countries (Woodruff et al., 2010). Initial 

13 results from this collaboration have examined PM and birth weight (Parker et al., 2011); 

14 work on O3 has not yet been performed. Although early animal studies (Kavlock et al., 

15 1980) found that exposure to O3 in the late gestation of pregnancy in rats led to some 

16 abnormal reproductive performances for neonates, to date human studies have reported 

17 inconsistent results for the association of ambient O3 on birth outcomes.

7.4.1 Effects on Sperm 

18 A limited amount of research has been conducted to examine the association between air 

19 pollution and male reproductive outcomes, specifically semen quality. To date, the 

20 epidemiologic studies have considered various exposure durations before semen 

21 collection that encompass either the entire period of spermatogenesis (i.e., 90 days) or 

22 key periods of sperm development that correspond to epididymal storage, development of 

23 sperm motility, and spermatogenesis. In an analysis conducted as part of the Teplice 

24 Program, 18-year-old men residing in the heavily polluted district of Teplice in the Czech 

25 Republic were found to be at greater risk of having abnormalities in sperm morphology 

26 and chromatin integrity than men of similar age residing in Prachatice, a less polluted 

27 district (Selevan et al., 2000; Sram et al., 1999). A follow-up longitudinal study 

28 conducted on a subset of the same men from Teplice revealed associations between total 

29 episodic air pollution and abnormalities in sperm chromatin (Rubes et al., 2005). A 

30 limitation of these studies is that they did not identify specific pollutants and their 

31 concentrations.

32 More recent epidemiologic studies conducted in the U.S. have also reported associations 

33 between ambient air pollution and sperm quality for individual air pollutants, including 

34 O3 and PM2.5. In a repeated measures study in Los Angeles, CA, Sokol et al. (2006)

35 reported a reduction in average sperm concentration during three exposure windows (0-9, 
Draft – Do Not Cite or Quote 7-39 September 2011

1 10-14, and 70-90 days before semen collection) associated with high ambient levels of 

2 O3 in healthy sperm donors. This effect persisted under a joint additive model for O3, 

3 CO, NO2 and PM10. The authors did not detect a reduction in sperm count. Hansen et al. 

4 (2010) investigated the effect of exposure to O3 and PM2.5 (using the same exposure 

5 windows used by Sokol et al. (2006) on sperm quality in three southeastern counties 

6 (Wake County, NC; Shelby County, TN; Galveston County, TX). Outcomes included 

7 sperm concentration and count, morphology, DNA integrity and chromatin maturity. 

8 Overall, the authors found both protective and adverse effects, although some results 

9 suggested adverse effects on sperm concentration, count and morphology. 

10 The biological mechanisms linking ambient air pollution to decreased sperm quality have 

11 yet to be determined, though O3-induced oxidative stress, inflammatory reactions, and 

12 the induction of the formation of circulating toxic species have been suggested as 

13 possible mechanisms (see Section 5. 3.8). Decremental effects on testicular morphology 

14 have been demonstrated in toxicological studies with histological evidence of O3-induced 

15 depletion of germ cells in testicular tissue and decreased seminiferous tubule epithelial 

16 layer. Jedlinska-Krakowska et al. (2006) demonstrated histopathological evidence of 

17 impaired spermatogenesis (round spermatids/ spermatocytes, giant spermatid cells, and 

18 focal epithelial desquamation with denudation to the basement membrane). The exposure 

19 protocol used five month old adult rats exposed to O3 as adults (0.5 ppm, 5 h/day for 

20 50 days). This degeneration could be rescued by vitamin E administration, indicating an 

21 antioxidant effect. Vitamin C administration had no effect at low doses of ascorbic acid 

22 and exacerbated the O3-dependent damage at high doses, as would be expected as 

23 vitamin C can be a radical generator instead of an antioxidant at higher doses. In 

24 summary, this study provided toxicological evidence of impaired spermatogenesis with 

25 O3 exposure that was rescued with certain antioxidant supplementation. 

26 Overall, there is limited epidemiologic evidence for an association with O3 concentration 

27 and decreased sperm concentration. A recent toxicological study provides limited 

28 evidence for a possible biological mechanism (histopathology showing impaired 

29 spermatogenesis) for such an association.

7.4.2 Effects on Reproduction

30 Evidence suggests that exposure to air pollutants during pregnancy is associated with 

31 adverse birth outcomes, which has been attributed to the increased susceptibility of the 

32 fetus due to physiologic immaturity. Gametes (i.e., ova and sperm) may be even more 

33 susceptible, especially outside of the human body, as occurs with assisted reproduction. 

34 Smokers require twice the number of in vitro fertilization (IVF) attempts to conceive as 
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1 non-smokers (Feichtinger et al., 1997), suggesting that a preconception exposure can be 

2 harmful to pregnancy. A recent study used an established national-scale, log-normal 

3 kriging method to spatially estimate daily mean concentrations of criteria pollutants at 

4 addresses of women undergoing their first IVF cycle and at their IVF labs from 2000 to 

5 2007 in the northeastern U.S. (Legro et al., 2010). Increasing O3 concentration at the 

6 patient’s address was significantly associated with an increased chance of live birth 

7 during ovulation induction (OR=1.13, [95% CI: 1.05, 1.22] per 10 ppb increase), but with 

8 decreased odds of live birth when exposed from embryo transfer to live birth (OR=0.79, 

9 [95% CI: 0.69, 0.90] per 10 ppb increase). After controlling for NO2 in a copollutant

10 model, however, O3 was no longer significantly associated with IVF failure. The results 

11 of this study suggest that exposure to O3 during ovulation was beneficial (perhaps due to 

12 early conditioning to O3), whereas later exposure to O3 (e.g., during gestation) was 

13 detrimental, and reduced the likelihood of a live birth.

14 In most toxicological studies, reproductive success appears to be unaffected by O3

15 exposure. Nonetheless, one study has reported that 25% of the BALB/c mouse dams in 

16 the highest O3 exposure group (1.2 ppm, GD9-18) did not complete a successful 

17 pregnancy, a significant reduction (Sharkhuu et al., 2011). Ozone administration 

18 (continuous 0.4, 0.8 or 1.2 ppm O3) to CD-1 mouse dams during the majority of 

19 pregnancy (PD7-17, which excludes the pre-implantation period), led to no adverse 

20 effects on reproductive success (proportion of successful pregnancies, litter size, sex 

21 ratio, frequency of still birth, or neonatal mortality) (Bignami et al., 1994). There was a 

22 nearly statistically significant increase in pregnancy duration (0.8 and 1.2 ppm O3). 

23 Initially, dam body weight (0.8 and 1.2 ppm), water consumption (0.4, 0.8 and 1.2 ppm 

24 O3) and food consumption (0.4, 0.8 and 1.2 ppm) during pregnancy were decreased with 

25 O3 exposure but these deficits dissipated a week or two after the initial exposure 

26 (Bignami et al., 1994). The anorexigenic effect of O3 exposure on the pregnant dam 

27 appears to dissipate with time; the dams seem to adapt to the O3 exposure. In males, data 

28 exist showing morphological evidence of altered spermatogenesis in O3 exposed animals 

29 (Jedlinska-Krakowska et al., 2006). Some evidence suggests that O3 may affect 

30 reproductive success when combined with other chemicals. Kavlock et al. (1979) showed 

31 that O3 acted synergistically with sodium salicylate to increase the rate of pup resorptions 

32 after midgestational exposure (1.0 ppm O3, GD9-12). At low doses of O3 exposure, 

33 toxicological studies show reproductive effects to include a transient anorexigenic effect 

34 of O3 on gestational weight gain, and a synergistic effect of O3 on salicylate-induced pup 

35 resorptions; other fecundity, pregnancy and gestation related outcomes appear unaffected 

36 by O3 exposure. 
Draft – Do Not Cite or Quote 7-41 September 2011

1 Collectively, there is very little epidemiologic evidence for the effect of O3 on 

2 reproductive success, and the reproductive success in rats appears to be unaffected in 

3 toxicological studies of O3 exposure. 

7.4.3 Birth Weight

4 With birth weight routinely collected in vital statistics and being a powerful predictor of 

5 infant mortality, it is the most studied outcome within air pollution-birth outcome 

6 research. Air pollution researchers have analyzed birth weight as a continuous variable 

7 and/or as a dichotomized variable in the form of LBW (<2,500 g [5 lbs, 8 oz]).

8 Birth weight is primarily determined by gestational age and intrauterine growth, but also 

9 depends on maternal, placental and fetal factors as well as on environmental influences. 

10 In both developed and developing countries, LBW is the most important predictor for 

11 neonatal mortality and is a significant determinant of postneonatal mortality and 

12 morbidity. Studies report that infants who are smallest at birth have a higher incidence of 

13 diseases and disabilities, which continue into adulthood (Hack and Fanaroff, 1999). 

14 The strongest evidence for an effect of O3 on birth weight comes from the Children’s 

15 Health Study conducted in southern California. In this study, Salam et al. (2005) report 

16 that maternal exposure to 24-h avg O3 concentrations averaged over the entire pregnancy 

17 was associated with reduced birth weight (39.3 g decrease [95% CI: -55.8, -22.8] in birth 

18 weight per 10 ppb and 8-h avg (19.2-g decrease [95% CI: -27.7, -10.7] in birth weight per 

19 10 ppb). This effect was stronger for concentrations averaged over the second and third 

20 trimesters. PM10, NO2 and CO concentrations averaged over the entire pregnancy were 

21 not statistically significantly associated with birth weight, although CO concentrations in 

22 the first trimester and PM10 concentrations in the third trimester were associated with a 

23 decrease in birth weight. Additionally, the authors observed a concentration-response 

24 relationship of birth weight with 24-h avg O3 concentrations averaged over the entire 

25 pregnancy that was clearest above the 30-ppb level (see Figure 7-4). Relative to the 

26 lowest decile of 24-h avg O3, estimates for the next 5 lowest deciles were approximately 

27 -40 g to -50 g, with no clear trend and with 95% confidence bounds that included zero. 

28 The highest four deciles of O3 exposure showed an approximately linear decrease in birth 

29 weight, and all four 95% CIs excluded zero, and ranged from mean decreases of 

30 74 grams to decreases of 148 grams. 
Draft – Do Not Cite or Quote 7-42 September 2011

Source: Salam et al. (2005)

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.

Figure 7-4 Birthweight deficit by decile of 24-h avg O3 concentration averaged 

over the entire pregnancy compared with the decile group with the 

lowest O3 exposure. 

1 Several additional studies conducted in the U.S. and Canada also investigated the 

2 association between ambient O3 concentrations and birth weight and report some weak 

3 evidence for an association. Morello-Frosch et al. (2010) estimated ambient O3

4 concentrations throughout pregnancy and for each trimester in the neighborhoods of 

5 women who delivered term singleton births between 1996 and 2006 in California. A 10-

6 ppb increase in O3 averaged across the entire pregnancy was associated with a 5.7-g 

7 decrease (95% CI: -6.6, -4.9) in birth weight when exposures were calculated using 

8 monitors within 10 km of the maternal address at date of birth. When the distance from 

9 the monitor was restricted to 3 km, the decrease in birth weight associated with a 10-ppb 

10 increase in O3 increased to 8.9 g (95% CI: -10.6, -7.1). These results persisted in 

11 copollutant models and in models that stratified by trimester of exposure, SES, and race. 

12 Darrow et al. (2011a) did not observe an association with birth weight and O3

13 concentrations during two exposure periods of interest (i.e., the first month and last 

14 trimester), but did find an association with reduced birth weight when examining the 

15 cumulative air pollution concentration during the entire pregnancy period. Additionally, 

16 they observed effect modification by race and ethnicity, such that associations between 

17 birth weight and third-trimester O3 concentrations were significantly stronger in 
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1 Hispanics and non-Hispanic African Americans than in non-Hispanic whites. Chen et al. 

2 (2002) used 8-h avg O3 concentrations to create exposure variables based on average 

3 maternal exposure for each trimester. Ozone was not found to be related to birth weight 

4 in single-pollutant models, though the O3 effect during the third trimester was borderline 

5 statistically significant in a copollutant model with PM10.

6 Several studies found no association between ambient O3 concentrations and birth 

7 weight. Wilhelm and Ritz (2005) extended previous analyses of term LBW (Ritz et al., 

8 2000; Ritz and Yu, 1999) to include the period 1994-2000. The authors examined varying 

9 residential distances from monitoring stations to see if the distance affected risk 

10 estimation, exploring the possibility that effect attenuation may result from local pollutant 

11 heterogeneity inadequately captured by ambient monitors. As in their previous studies, 

12 the authors observed associations between elevated concentrations of CO and PM10 both 

13 early and late in pregnancy and risk of term LBW. After adjusting for CO and/or PM10

14 the authors did not observe associations between O3 and term LBW in any of their 

15 models. Brauer et al. (2008) evaluated the impacts of air pollution (CO, NO2, NO, O3, 

16 SO2, PM2.5, PM10) on birth weight for the period 1999-2002 using spatiotemporal 

17 residential exposure metrics by month of pregnancy in Vancouver, BC. Quantitative 

18 results were not presented for the association between O3 and LBW, though the authors 

19 observed associations that were largely protective. Dugandzic et al. (2006) examined the 

20 association between LBW and ambient levels of air pollutants by trimester of exposure 

21 among a cohort of term singleton births from 1988-2000. Though there was some 

22 indication of an association with SO2 and PM10, there were no effects for O3.

23 Similarly, studies conducted in Australia, Latin America, and Asia report limited 

24 evidence for an association between ambient O3 and measures of birth weight. In Sydney, 

25 Australia, Mannes et al. (2005) found that O3 concentrations in the second trimester of 

26 pregnancy had small adverse effects on birth weight (7.5-g decrease; [95 % CI: -13.8, 

27 1.2] per 10 ppb), although this effect disappeared when the analysis was limited to births 

28 with a maternal address within 5 km of a monitoring station (87.7-g increase; [95% CI: 

29 10.5, 164.9] per 10 ppb). Hansen et al. (2007) reported that trimester and monthly 

30 specific exposures to all pollutants were not statistically significantly associated with a 

31 reduction in birth weight in Brisbane, Australia. In Sao Paulo, Brazil, Gouveia et al. 

32 (2004) found that O3 exhibited a small inverse relation with birth weight over the third 

33 trimester (6.0-g decrease; [95% CI: -30.8, 18.8] per 10 ppb). Lin et al. (2004b) reported a 

34 positive, though not statistically significant, exposure-response relationship for O3 during 

35 the entire pregnancy in a Taiwanese study. In a study performed in Korea, Ha et al. 

36 (2001) reported no O3 effect during the first trimester of pregnancy, but they found that 

37 during the third trimester of pregnancy O3 was associated with LBW (RR=1.05 [95% CI: 

38 1.02, 1.08] per 10 ppb). 
Draft – Do Not Cite or Quote 7-44 September 2011

Table 7-5 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. (2011a) 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, US

(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

Mannes et al. (2005) Sydney, Australia

(n=138,056)

1-h max: 

31.6

City-wide 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

City-wide 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

City-wide 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 City-wide 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 7-5 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 
Draft – Do Not Cite or Quote 7-45 September 2011

1 results of the Children’s Health Study. Additional studies conducted in the U.S., Canada, 

2 Australia, Latin America, and Asia provide limited and inconsistent evidence to support 

3 the effect reported in the Children’s Health Study. The toxicological literature on the 

4 effect of O3 on birth weight is sparse. In some studies, the reporting of birth weight may 

5 be avoided because birth weight can be confounded by decreased litter size resulting 

6 from an increased rate of pup resorption (aborted pups) in O3 exposed dams. In one 

7 toxicological study by Haro and Paz (1993), no differences in litter size were observed 

8 and decreased birth weight in pups from dams who were exposed to 1ppm O3 during 

9 pregnancy was reported. A second animal toxicology study recapitulated these finding 

10 with pregnant BALB/c mice that exposed to O3 (1.2 ppm, GD9-18) producing pups with 

11 significantly decreased birth weights (Sharkhuu et al., 2011).

7.4.4 Preterm Birth 

12 Preterm birth (PTB) is a syndrome (Romero et al., 2006) that is characterized by multiple 

13 etiologies. It is therefore unusual to be able to identify an exact cause for each PTB. In 

14 addition, PTB is not an adverse outcome in itself, but an important determinant of health 

15 status (i.e., neonatal morbidity and mortality). Although some overlap exists for common 

16 risk factors, different etiologic entities related to distinct risk factor profiles and leading 

17 to different neonatal and postneonatal complications are attributed to PTB and measures 

18 of fetal growth. Although both restricted fetal growth and PTB can result in LBW, 

19 prematurity does not have to result in LBW or growth restricted babies. 

20 A major issue in studying environmental exposures and PTB is selecting the relevant 

21 exposure period, since the biological mechanisms leading to PTB and the critical periods 

22 of vulnerability are poorly understood (Bobak, 2000). Exposures proximate to the birth 

23 may be most relevant if exposure causes an acute effect. However, exposure occurring in 

24 early gestation might affect placentation, with results observable later in pregnancy, or 

25 cumulative exposure during pregnancy may be the most important determinant. The 

26 studies reviewed have dealt with this issue in different ways. Many have considered 

27 several exposure metrics based on different periods of exposure. Often the time periods 

28 used are the first month (or first trimester) of pregnancy and the last month (or 6 weeks) 

29 prior to delivery. Using a time interval prior to delivery introduces an additional problem 

30 since cases and controls are not in the same stage of development when they are 

31 compared. For example, a preterm infant delivered at 36 weeks is a 32-week fetus 

32 4 weeks prior to birth, while an infant born at term (40 weeks) is a 36-week fetus 4 weeks 

33 prior to birth.
Draft – Do Not Cite or Quote 7-46 September 2011

1 Recently, investigators have examined the association of PTB with both short-term (i.e., 

2 hours, days, or weeks) and long-term (i.e., months or years) exposure periods. Time3 series studies have been used to examine the association between air pollution 

4 concentrations during the days immediately preceding birth. An advantage of these time5 series studies is that this approach can remove the influence of covariates that vary across 

6 individuals over a short period of time. Retrospective cohort and case-control studies 

7 have been used to examine long-term exposure periods, often averaging air pollution 

8 concentrations over months or trimesters of pregnancy.

9 Reported studies fail to show consistency in pollutants and periods during pregnancy 

10 when an effect occurs. For example, while some studies find the strongest effects 

11 associated with exposures early in pregnancy, others report effects when the exposure is 

12 limited to the second or third trimester. However, the effect of air pollutant exposure 

13 during pregnancy on PTB has a biological basis. There is an expanding list of possible 

14 mechanisms that may explain the association between O3 exposure and PTB (see 

15 Section 5. 4.2.4).

16 Many studies of PTB compare exposure in quartiles, using the lowest quartile as the 

17 reference (or control) group. No studies use a truly unexposed control group. If exposure 

18 in the lowest quartile confers risk, than it may be difficult to demonstrate additional risk 

19 associated with a higher quartile. Thus negative studies must be interpreted with caution.

20 Preterm birth occurs both naturally (idiopathic preterm), and as a result of medical 

21 intervention (iatrogenic preterm). Ritz et al. (2007; 2000) excluded all births by Cesarean 

22 section to limit their studies to idiopathic preterm. No other studies attempted to 

23 distinguish the type of PTB, although air pollution exposure maybe associated with only 

24 one type. This is a source of potential effect misclassification. 

25 Generally, studies of air pollution-birth outcome conducted in North America and the 

26 United Kingdom have not identified an association between PTB and maternal exposure 

27 to O3. Most recently, Darrow et al. (2009) used vital record data to construct a 

28 retrospective cohort of 476,489 births occurring between 1994 and 2004 in 5 central 

29 counties of metropolitan Atlanta. Using a time-series approach, the authors examined 

30 aggregated daily counts of PTB in relation to ambient levels of CO, NO2, SO2, O3, PM10, 

31 PM2.5 and speciated PM measurements. This study investigated 3 gestational windows of 

32 exposure: the first month of gestation, the final week of gestation, and the final 6 weeks 

33 of gestation. The authors did not observe associations of PTB with O3.

34 A number of U.S. studies were conducted in southern California, and report somewhat 

35 inconsistent results. Ritz et al. (2000) evaluated the effect of air pollution (CO, NO2, O3, 

36 PM10) exposure during pregnancy on the occurrence of PTB in a cohort of 97,518 
Draft – Do Not Cite or Quote 7-47 September 2011

1 neonates born in southern California between 1989 and 1993. The authors use both short2 and long-term exposure windows, averaging pollutant measures taken at the closest air3 monitoring station over distinct periods, such as 1, 2, 4, 6, 8, 12, and 26 weeks before 

4 birth and the whole pregnancy period. Additionally, they calculated average exposures 

5 for the first and second months of pregnancy. The authors found no consistent effects for 

6 O3 over any of the pregnancy periods in single or multi-pollutant models. Wilhelm and 

7 Ritz (2005) extended previous analyses of PTB (Ritz et al., 2000; Ritz and Yu, 1999) in 

8 California to include 1994-2000. The authors examined varying residential distances 

9 from monitoring stations to see if the distance affected risk estimation, because effect 

10 attenuation may result from local pollutant heterogeneity inadequately captured by 

11 ambient monitors. The authors analyzed the association between O3 exposure during 

12 varying periods of pregnancy and PTB, finding a positive association between O3 levels 

13 in both the first trimester of pregnancy (RR=1.23 [95% CI: 1.06, 1.42] per 10 ppb 

14 increase in 24-h avg O3) and the first month of pregnancy (results for first trimester 

15 exposure were similar, but slightly smaller, quantitative results not presented) in models 

16 containing all pollutants. No association was observed between O3 in the 6 weeks before 

17 birth and preterm delivery. Finally, Ritz et al. (2007) conducted a case-control survey 

18 nested within a birth cohort and assessed the extent to which residual confounding and 

19 exposure misclassification impacted air pollution effect estimates. The authors calculated 

20 mean exposure levels for three gestational periods: the entire pregnancy, the first 

21 trimester, and the last 6 weeks before delivery. Though positive associations were 

22 observed for CO and PM2.5, no consistent patterns of increase in the odds of PTB for O3

23 or NO2 were observed. 

24 One study conducted in Canada evaluated the impacts of air pollution (including CO, 

25 NO2, NO, O3, SO2, PM2.5, and PM10) on PTBs (1999-2002) using spatiotemporal 

26 residential exposure metrics by month of pregnancy in Vancouver, BC (Brauer et al., 

27 2008). The authors did not observe consistent associations with any of the pregnancy 

28 average exposure metrics except for PM2.5 for PTB. The O3 associations were largely 

29 protective, and no quantitative results were presented for O3. Additionally, Lee et al. 

30 (2008c) used time-series techniques to investigate the short-term associations of O3 and 

31 PTB in London, England. In addition to exposure on the day of birth, cumulative 

32 exposure up to 1 week before birth was investigated. The risk of PTB did not increase 

33 with exposure to the levels of ambient air pollution experienced by this population. 

34 Conversely, studies conducted in Australia and China provide evidence for an association 

35 between ambient O3 and PTB. Hansen et al. (2006) reported that exposure to O3 during 

36 the first trimester was associated with an increased risk of PTB (OR=1.38, [95% CI: 

37 1.14, 1.69] per 10 ppb increase). Although the test for trend was significant due to the 

38 strong effect in the highest quartile, there was not an obvious exposure-response pattern 
Draft – Do Not Cite or Quote 7-48 September 2011

1 across the quartiles of O3 during the first trimester. The effect estimate was diminished 

2 and lost statistical significance when PM10 was included in the model (OR=1.23, [95% 

3 CI: 0.97, 1.59] per 10 ppb increase). Maternal exposure to O3 during the 90 days prior to 

4 birth showed a weak, positive association with PTB (OR=1.09, [95% CI: 0.85, 1.39] per 

5 10 ppb increase). Jalaludin et al. (2007) found that O3 levels in the month and 

6 three months preceding birth had a statistically significant association with PTB. Ozone 

7 levels in the first trimester of pregnancy were associated with increased risks for PTBs 

8 (OR=1.15 [95% CI: 1.05, 1.24] per 10 ppb increase in 1-h max O3 concentration), and 

9 remained a significant predictor of PTB in copollutant models (ORs between 1.07 and 

10 1.10). ORs increased for first month of pregnancy when restricted to within 5 km of a 

11 monitoring station (OR=1.60, [95% CI: 1.27, 2.03]), but did not show a cumulative effect 

12 for first 3 months of pregnancy (OR=0.81, [95% CI: 0.67, 0.98]). Jiang et al. (2007)

13 examined the acute effect of air pollution on PTB, including risk in relation to levels of 

14 pollutants for a single day exposure window with lags from 0 to 6 days before birth. An 

15 increase of 10 ppb of the 8-week avg of O3 corresponded to 9.47 % (95% CI: 0.70, 

16 18.7%) increase in PTBs. Increases in PTB were also observed for PM10, SO2, and NO2. 

17 The authors did not observe any significant acute effect of outdoor air pollution on PTB 

18 among the 1-day acute time windows examined in the week before birth. 

19 Little data is available from toxicological studies; one study reported a nearly statistically 

20 significant increase in pregnancy duration in mice when exposed to 0.8 or 1.2 ppm O3. 

21 This phenomenon was most likely due to the anorexigenic effect of relatively high O3

22 concentrations (Bignami et al., 1994).
Draft – Do Not Cite or Quote 7-49 September 2011

Table 7-6 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, US

(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

City-wide 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

City-wide avg and <5 

km from monitor

First month: 1.604 (1.268, 2.030)b

T1: 0.807 (0.668, 0.976)b

T3: 1.011 (0.910, 1.124)b

Last month: 0.984 (0.906, 1.069)b

Jiang et al. (2007) Shanghai, China

(n=3,346 preterm 

births)

8-h avg: 

32.7

City-wide 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.

b

Relative risk of PTB per 1 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

1 Table 7-6 provides a brief overview of the epidemiologic studies of PTB. In summary, 

2 the evidence is consistent when examining shorter-term, late-pregnancy exposure to O3

3 and reports no association with PTB. However when long-term exposure to O3 early in 

4 pregnancy is examined the results are inconsistent. Studies conducted in the U.S., 
Draft – Do Not Cite or Quote 7-50 September 2011

1 Canada, and England find no association with O3 and PTB, while studies conducted in 

2 Australia and China report an O3 effect on PTB.

7.4.5 Fetal Growth 

3 Low birth weight has often been used as an outcome measure because it is easily 

4 available and accurately recorded on birth certificates. However, LBW may result from 

5 either short gestation, or inadequate growth in utero. Most of the studies investigating air 

6 pollution exposure and LBW limited their analyses to term infants to focus on inadequate 

7 growth. A number of studies were identified that specifically addressed growth restriction 

8 in utero by identifying infants who failed to meet specific growth standards. Usually 

9 these infants had birth weight less than the 10th percentile for gestational age, using an 

10 external standard. Many of these studies have been previously discussed, since they also 

11 examined other reproductive outcomes (i.e., LBW or PTB).

12 Fetal growth is influenced by maternal, placental, and fetal factors. The biological 

13 mechanisms by which air pollutants may influence the developing fetus remain largely 

14 unknown. Several mechanisms have been proposed, and are the same as those 

15 hypothesized for birth weight (see Section 5. 4.2.4). Additionally, in animal toxicology 

16 studies, O3 causes transient anorexia in exposed pregnant dams. This may be one of 

17 many possible contributors to O3-dependent decreased fetal growth.

18 A limitation of environmental studies that use birth weight as a proxy measure of fetal 

19 growth is that patterns of fetal growth during pregnancy cannot be assessed. This is 

20 particularly important when investigating pollutant exposures during early pregnancy as 

21 birth weight is recorded many months after the exposure period. The insult of air 

22 pollution may have a transient effect on fetal growth, where growth is hindered at one 

23 point in time but catches up at a later point. For example, maternal smoking during 

24 pregnancy can alter the growth rate of individual body segments of the fetus at variable 

25 developmental stages, as the fetus experiences selective growth restriction and 

26 augmentation (Lampl and Jeanty, 2003).

27 The terms small-for-gestational-age (SGA), which is defined as a birth weight <10th 

28 percentile for gestational age (and often sex and/or race), and intrauterine growth 

29 retardation (IUGR) are often used interchangeably. However, this definition of SGA does 

30 have limitations. For example, using it for IUGR may overestimate the percentage of 

31 “growth-restricted” neonates as it is unlikely that 10% of neonates have growth 

32 restriction (Wollmann, 1998). On the other hand, when the 10th percentile is based on the 

33 distribution of live births at a population level, the percentage of SGA among PTB is 

34 most likely underestimated (Hutcheon and Platt, 2008). Nevertheless, SGA represents a 
Draft – Do Not Cite or Quote 7-51 September 2011

1 statistical description of a small neonate, whereas the term IUGR is reserved for those 

2 with clinical evidence of abnormal growth. Thus all IUGR neonates will be SGA, but not 

3 all SGA neonates with be IUGR (Wollmann, 1998). In the following section the terms 

4 SGA and IUGR are referred to as each cited study used the terms.

5 Over the past decade a number of studies examined various metrics of fetal growth 

6 restriction. Salam et al. (2005) assessed the effect of increasing O3 concentrations on 

7 IUGR in a population of infants born in California from 1975-1987 as part of the 

8 Children’s Health Study. The authors reported that maternal O3 exposures averaged over 

9 the entire pregnancy and during the third trimester were associated with increased risk of 

10 IUGR. A 10-ppb difference in 24-h maternal O3 exposure during the third trimester 

11 increased the risk of IUGR by 11% (95% CI: 0, 20%). Brauer et al. (2008) evaluated the 

12 impacts of air pollution (CO, NO2, NO, O3, SO2, PM2.5, PM10) on SGA (1999-2002) 

13 using spatiotemporal residential exposure metrics by month of pregnancy in Vancouver, 

14 BC. The O3 associations were largely protective (OR= 0.87, [95% CI: 0.81, 0.93] for a 

15 10 ppb increase in inverse distance weighted SGA), and no additional quantitative results 

16 were presented for O3. Liu et al. (2007b) examined the association between IUGR among 

17 singleton term live births and SO2, NO2, CO, O3, and PM2.5 in 3 Canadian cities for the 

18 period 1985-2000. No increase in the risk of IUGR in relation to exposure to O3 averaged 

19 over each month and trimester of pregnancy was noted. 

20 Three studies conducted in Australia provide evidence for an association between 

21 ambient O3 and fetal growth restriction. Hansen et al. (2007) examined SGA among 

22 singleton, full-term births in Brisbane, Australia in relation to ambient air pollution (bsp, 

23 PM10, NO2, O3) during pregnancy. They also examined head circumference and crown24 heel length in a subsample of term neonates. Trimester specific exposures to all pollutants 

25 were not statistically significantly associated with a reduction in head circumference or 

26 an increased risk of SGA. When monthly-specific exposures were examined, the authors 

27 observed an increased risk of SGA associated with exposure to O3 during month 4 

28 (OR=1.11 [95% CI: 1.00, 1.24] per 10 ppb increase). In a subsequent study, Hansen et al. 

29 (2008) examined the possible associations between fetal ultrasonic measurements and 

30 ambient air pollution (PM10, O3, NO2, SO2) during early pregnancy. This study had two 

31 strengths: (1) fetal growth was assessed during pregnancy as opposed to at birth; and (2) 

32 there was little delay between exposures and fetal growth measurements, which reduces 

33 potential confounding and uses exposures that are concurrent with the observed growth 

34 pattern of the fetus. Fetal ultrasound biometric measurements were recorded for biparietal 

35 diameter (BPD), femur length, abdominal circumference, and head circumference. To 

36 further improve exposure assessment, the authors restricted the samples to include only 

37 scans from women for whom the centroid of their postcode was within 14 km of an air 

38 pollution monitoring site. Ozone during days 31-60 was associated with decreases in all 
Draft – Do Not Cite or Quote 7-52 September 2011

1 of the fetal growth measurements, and a 1.78 mm reduction in abdomen circumference 

2 per 10 ppb increase in O3 concentration, though this effect did not persist in copollutant

3 models. The change in ultrasound measurements associated with O3 during days 31-60 of 

4 gestation indicated that increasing O3 concentration decreased the magnitude of 

5 ultrasound measurements for women living within 2 km of the monitoring site. The 

6 relationship decreased toward the null as the distance from the monitoring sites increased. 

7 When assessing effect modification due to SES, there was some evidence of effect 

8 modification for most of the associations, with the effects of air pollution stronger in the 

9 highest SES quartile. In the third study, Mannes et al. (2005) estimated the effects of 

10 pollutant (PM10, PM2.5, NO2, CO and O3) exposure in the first, second and third 

11 trimesters of pregnancy and risk of SGA in Sydney, Australia. Citywide average air 

12 pollutant concentrations in the last month, third trimester, and first trimester of pregnancy 

13 had no effect on SGA. Concentrations of O3 in the second trimester of pregnancy had 

14 small but adverse effects on SGA (OR=1.10 [95% CI: 1.00, 1.14] per 10 ppb increment). 

15 This effect disappeared when the analysis was limited to births with a maternal address 

16 within 5 km of a monitoring station (OR=1.00 [95% CI: 0.60, 1.79] per 10 ppb 

17 increment).

18 Very little information from toxicological studies is available to address effects on fetal 

19 growth. However, there is evidence to suggest that prenatal exposure to O3 can affect 

20 postnatal growth. A few studies reported that mice or rats exposed developmentally 

21 (gestationally ± lactationally) to O3 had deficits in body weight gain in the postpartum 

22 period (Bignami et al., 1994; Haro and Paz, 1993; Kavlock et al., 1980). 

23 Table 7-7 provides a brief overview of the epidemiologic studies of fetal growth 

24 restriction. In summary, the evidence is inconsistent when examining exposure to O3 and 

25 fetal growth restriction. Similar to PTB, studies conducted in Australia have reported an 

26 effect of O3 on fetal growth, whereas studies conducted in other areas have not found 

27 such an effect. This may be due to the restriction of births to those within 2-14 km of a 

28 monitoring station, as was done in the Australian studies. 
Draft – Do Not Cite or Quote 7-53 September 2011

Table 7-7 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

City-wide 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

City-wide 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.

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 Time-pregnant BALB/c mice were exposed to O3 (0, 0.4, 0.8, or 1.2 ppm) GD9-18 with 

2 parturition at GD20-21 (Sharkhuu et al., 2011). As the offspring aged, postnatal litter 

3 body weight continued to be significantly decreased in the highest dose (1.2 ppm) O3

4 group at PND3 and PND7. When the pups were weighed separately by sex at PND42, the 

5 males with the highest dose of O3 exposure (1.2 ppm, GD9-18) had significant 

6 decrements in body weight (Sharkhuu et al., 2011).
Draft – Do Not Cite or Quote 7-54 September 2011

7.4.7 Birth Defects 

1 Despite the growing body of literature evaluating the association between ambient air 

2 pollution and various adverse birth outcomes, relatively few studies have investigated the 

3 effect of temporal variations in ambient air pollution on birth defects. Heart defects and 

4 oral clefts have been the focus of the majority of these recent studies, given the higher 

5 prevalence than other birth defects and associated mortality. Mechanistically, air 

6 pollutants could be involved in the etiology of birth defects via a number of key events 

7 (see Section 5. 4.2.4). 

8 Several studies have been conducted examining the relationship between O3 exposure 

9 during pregnancy and birth defects and reported a positive association with cardiac 

10 defects. The earliest of these studies was conducted in southern California (Ritz et al., 

11 2002). This study evaluated the effect of air pollution on the occurrence of cardiac birth 

12 defects in neonates and fetuses delivered in southern California in 1987-1993. Maternal 

13 exposure estimates were based on data from the fixed site closest to the mother’s ZIP 

14 code area. When using a case-control design where cases were matched to 10 randomly 

15 selected controls, results showed increased risks for aortic artery and valve defects 

16 (OR=1.56 [95% CI: 1.16, 2.09] per 10 ppb O3), pulmonary artery and valve anomalies 

17 (OR=1.34 [95% CI: 0.96, 1.87] per 10 ppb O3), and conotruncal defects (OR=1.36 [95% 

18 CI: 0.91, 2.03] per 10 ppb O3) in a dose-response manner with second-month O3

19 exposure. A study conducted in Texas (Gilboa et al., 2005) looked at a similar period of 

20 exposure but reported no association with most of the birth defects studied (O3

21 concentration was studied using quartiles with the lowest representing <18 ppb and the 

22 highest representing ≥ 31 ppb). The authors found slightly elevated odds ratios for 

23 pulmonary artery and valve defects. They also detected an inverse association between 

24 O3 exposure and isolated ventricular septal defects. Overall, this study provided some 

25 weak evidence that air pollution increases the risk of cardiac defects. Hansen et al. (2009)

26 investigated the possible association between ambient air pollution and the risk of cardiac 

27 defects. When analyzing all births with exposure estimates for O3 from the nearest 

28 monitor there was no indication for an association with cardiac defects. There was also 

29 no adverse association when restricting the analyses to only include births where the 

30 mother resided within 12 km of a monitoring station. However, among births within 6 km 

31 of a monitor, a 10 ppb increase in O3 was associated with an increased risk of pulmonary 

32 artery and valve defects (OR=8.76 [95% CI: 1.80, 56.55]). As indicated by the very wide 

33 credible intervals, there were very few cases in the sensitivity analyses for births within 6 

34 km of a monitor, and this effect could be a result of type I errors. Dadvand et al. (2011)

35 investigated the association between maternal exposure to ambient air pollution and the 

36 occurrence of cardiac birth defects in England. Similar to Hansen et al. (2009), they 

37 found no associations with maternal exposure to O3 except for when the analysis was 
Draft – Do Not Cite or Quote 7-55 September 2011

1 limited to those subjects residing within a 16 km distance of a monitoring station (OR for 

2 malformations of pulmonary and tricuspid valves=1.64 [95% CI: 1.04, 2.60] per 10 ppb 

3 increase in O3).

4 Despite the association between O3 and cardiac defects observed in the above studies, a 

5 recent study did not observe an increased risk of cardiac birth defects associated with 

6 ambient O3 concentrations. The study, conducted in Atlanta, GA, examined O3 exposure 

7 during the third through seventh week of pregnancy and reported no association with risk 

8 of cardiovascular malformations (mean long-term average of 8-h O3 concentrations 

9 excluding November through February ranged by 5-year groups from 39.8 to 43.3 ppb) 

10 (Strickland et al., 2009). 

11 Several of these studies have also examined the relationship between O3 exposure during 

12 pregnancy and oral cleft defects. The study by Ritz et al. (Ritz et al., 2002) evaluated the 

13 effect of air pollution on the occurrence of orofacial birth defects and did not observe 

14 strong associations between ambient O3 concentration and orofacial defects. They did 

15 report an OR of 1.13 (95% CI: 0.90, 1.40) per 10 ppb during the second trimester for cleft 

16 lip with or without cleft palate. Similarly, Gilboa et al. (Gilboa et al., 2005) reported and 

17 OR of 1.09 (95% CI: 0.70, 1.69) for oral cleft defects when the fourth quartile was 

18 contrasted with the first quartile of exposure during 3-8 weeks of pregnancy. Hansen et 

19 al. (2009) reported no indication for an association with cleft defects. Hwang and Jaakola 

20 (2008) conducted a population-based case-control study to investigate exposure to 

21 ambient air pollution and the risk of cleft lip with or without cleft palate in Taiwan. The 

22 risk of cleft lip with or without cleft palate was increased in relation to O3 levels in the 

23 first gestational month (OR=1.17 [95% CI: 1.01, 1.36] per 10 ppb) and second gestational 

24 month (OR=1.22 [95% CI: 1.03, 1.46] per 10 ppb), but was not related to any of the other 

25 pollutants. In three-pollutant models, the effect estimates for O3 exposure were stable for 

26 the four different combinations of pollutants and were all statistically significant.

27 Marshall et al. (2010) compared estimated exposure to ambient pollutants during early 

28 pregnancy among mothers of children with oral cleft defects to that among mothers of 

29 controls. The authors observed no consistent elevated associations between any of the air 

30 pollutants examined and cleft malformations, though there was a weak association 

31 between cases of cleft palate only and increasing O3 concentrations. This association 

32 increased when cases and controls were limited to those with residences within 10 km of 

33 the closest O3 monitor (OR=2.2 [95% CI: 1.0, 4.9], comparing highest quartile [>33 ppb] 

34 to lowest quartile [<15 ppb]). 

35 A limited number of toxicological studies have examined birth defects in animals 

36 exposed gestationally to O3. Kavlock et al. (1979) exposed pregnant rats to O3 for precise 

37 periods during organogenesis. No significant teratogenic effects were found in rats 
Draft – Do Not Cite or Quote 7-56 September 2011

1 exposed 8 hr/day to concentrations of O3 varying from 0.44 to 1.97 ppm during early 

2 (days 6-9), mid (days 9-12), or late (days 17 to 20) gestation, or the entire period of 

3 organogenesis (days 6-15). Earlier research found eyelid malformation following 

4 gestational and postnatal exposure to 0.2 ppm O3 (Veninga, 1967). 

5 Table 7-8 provides a brief overview of the epidemiologic studies of birth defects. These 

6 studies have focused on cardiac and oral cleft defects, and the results from these studies 

7 are not entirely consistent. This inconsistency could be due to the absence of true 

8 associations between O3 and risks of cardiovascular malformations and oral cleft defects; 

9 it could also be due to differences in populations, pollution levels, outcome definitions, or 

10 analytical approaches. The lack of consistency of associations between O3 and 

11 cardiovascular malformations or oral cleft defects might be due to issues relating to 

12 statistical power or measurement error. A recent meta-analysis of air pollution and 

13 congenital anomalies concluded that there was no statistically significant increase in risk 

14 of congenital anomalies and O3 (Vrijheid et al., 2011). These authors note that 

15 heterogeneity in the results of these studies may be due to inherent differences in study 

16 location, study design, and/or analytic methods, and comment that these studies have not 

17 employed some recent advances in exposure assessment used in other areas of air 

18 pollution research that may help refine or reduce this heterogeneity.

Table 7-8 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 Jaakola 

(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 City-wide 

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
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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 Sections 7.2.1 and 7.2.3, and are 

8 included here because they reported both prenatal and post-natal exposure periods. 

9 Mortimer et al. (2008a, b) examined the association of prenatal and lifetime exposures to 

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

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

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

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

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

15 entire lifetime. The 8-h avg O3 concentrations were approximately 50 ppb for each of the 

16 exposure metrics (estimated from figure). In the first analysis (Mortimer et al., 2008a), 

17 negative effects on pulmonary function were found for exposure to PM10, NO2, and CO 

18 during key neonatal and early life developmental periods. The authors did not find a 

19 negative effect of exposure to O3 among this cohort. In the second analysis (Mortimer et 

20 al., 2008b), sensitization to at least one allergen was associated, in general, with higher 

21 levels of CO and PM10 during the entire pregnancy and second trimester and higher PM10

22 during the first 2 years of life. Lower exposure to O3 during the entire pregnancy or 

23 second trimester was associated with an increased risk of allergen sensitization. Although 

24 the pollutant metrics across time periods are correlated, the strongest associations with 

25 the outcomes were observed for prenatal exposures. Though it may be difficult to 

26 disentangle the effect of prenatal and postnatal exposures, the models from this group of 

27 studies suggest that each time period of exposure may contribute independently to 

28 different dimensions of school-aged children’s pulmonary function. For 4 of the 8 

29 pulmonary-function measures (FVC, FEV1, PEF, FEF25-75), prenatal exposures were 

30 more influential on pulmonary function than early-lifetime metrics, while, in contrast, the 

31 ratio of measures (FEV1/FVC and FEF25-75/FVC) were most influenced by postnatal 

32 exposures. When lifetime metrics were considered alone, or in combination with the 

33 prenatal metrics, the lifetime measures were not associated with any of the outcomes, 

34 suggesting the timing of the exposure may be more important than the overall dose and 

35 prenatal exposures are not just markers for lifetime or current exposures. 
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1 Clark et al. (2010) investigated the effect of exposure to ambient air pollution in utero 

2 and during the first year of life on risk of subsequent asthma diagnosis (incident asthma 

3 diagnosis up to age 3-4) in a population-based nested case-control study. Air pollution 

4 exposure for each subject based on their residential address history was estimated using 

5 regulatory monitoring data, land use regression modeling, and proximity to stationary 

6 pollution sources. An average exposure was calculated for the duration of pregnancy 

7 (~15 ppb) and the first year of life (~14 ppb). In contrast to the Mortimer et al. studies 

8 (2008a, b), the effect estimates for first-year exposure were generally larger than for in 

9 utero exposures. However, similar to the Mortimer et al. studies, the observed 

10 associations with O3 were largely protective. Because of the relatively high correlation 

11 between in utero and first-year exposures for many pollutants, it was difficult to discern 

12 the relative importance of the individual exposure periods. 

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

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

15 inhomogeneity and eNO were measured in 241 unsedated, sleeping neonates (age= 

16 5 weeks). The median of the 24-h avg O3 concentrations averaged across the post-natal 

17 period was ~44 ppb. Consistent with the previous studies, no association was found for 

18 prenatal exposure to O3 and lung function.

19 The new toxicological literature since the 2006 O3 AQCD, covering respiratory changes 

20 related to developmental O3 exposure, reports ultrastructural changes in bronchiole 

21 development, alterations in placental and pup cytokines, and increased pup airway hyper22 reactivity. These studies are detailed below. Older studies are discussed where new 

23 information is not available. 

24 Fetal rat lung bronchiole development is triphasic, comprised of the glandular phase 

25 (measured at GD18), the canalicular phase (GD20), and the saccular phase (GD21). The 

26 ultrastructural lung development in fetuses of pregnant rats exposed to 1-ppm O3 (12 

27 h/day, out to either GD18, GD20 or GD21) was examined by electron microscopy during 

28 these three phases. In the glandular phase, bronchiolar columnar epithelial cells in fetuses 

29 of dams exposed to O3 had cytoplasmic damage and swollen mitochondria. Bronchial 

30 epithelium at the canalicular phase in O3 exposed pups had delayed maturation in 

31 differentiation, i.e., glycogen abundance in secretory cells had not diminished as it should 

32 with this phase of development. Congruent with this finding, delayed maturation of 

33 tracheal epithelium following early neonatal O3 exposure (1 ppm, 4-5 h/day for 

34 first week of life) in lambs has been previously reported (Mariassy et al., 1990; Mariassy 

35 et al., 1989). Also at the canalicular phase, atypical cells were seen in the bronchiolar 

36 lumen of O3 exposed rat fetuses. Finally, in the saccular phase, mitochondrial 

37 degradation was present in the non-ciliated bronchiolar cells of rats exposed in utero to 
Draft – Do Not Cite or Quote 7-59 September 2011

1 O3. In conclusion, O3 exposure of pregnant rats produced ultra-structural damage to near2 term fetal bronchiolar epithelium (López et al., 2008). 

3 Exposure of laboratory animals to multiple airborne pollutants can differentially affect 

4 pup physiology. One study showed that exposure of C57BL/6 mouse dams to 0.48 mg 

5 PM intratracheally twice weekly for 3 weeks during pregnancy augmented O3-induced 

6 airway hyper-reactivity in juvenile offspring. Maternal PM exposure also significantly 

7 increased placental cytokines above vehicle-instilled controls. Pup postnatal O3 exposure 

8 (1 ppm 3 h/day, every other day, thrice weekly for 4 weeks) induced significantly 

9 increased cytokine levels (IL-1β, TNF-α, KC, and IL-6) in whole lung versus postnatal 

10 air exposed groups; this was further exacerbated with gestational PM exposure (Auten et 

11 al., 2009). 

12 A series of experiments using infant rhesus monkeys repeatedly exposed to 0.5 ppm O3

13 starting at one-month of age have examined the effect of O3 alone or in combination with 

14 an inhaled allergen on morphology and lung function (Plopper et al., 2007). Exposure to 

15 O3 alone or allergen alone produced small but not statistically significant changes in 

16 baseline airway resistance and airway responsiveness, but the combined exposure to both 

17 O3 + antigen produced statistically significant and greater than additive changes in both 

18 functional measurements. Additionally, cellular changes and significant structural 

19 changes in the respiratory tract have been observed in infant rhesus monkeys exposed to 

20 O3 (Fanucchi et al., 2006). A more detailed description of these studies can be found in 

21 Section 7.2.3 (Pulmonary Structure and Function), with mechanistic information found in 

22 Section 5.4.2.4.

23 Lung immunological response in O3 exposed pups was followed by analyzing BAL and 

24 lung tissue. Sprague Dawley (SD) pups were exposed to a single 3h exposure of air or O3

25 (0.6 ppm) on PND 13 (Han et al., 2011). Bronchoalveolar lavage (BAL) was performed 

26 10 hours after the end of O3 exposure. BALF polymorphonuclear leukocytes (PMNs) and 

27 total BALF protein were significantly elevated in O3 exposed pups. Lung tissue from O3

28 exposed pups had significant elevations of manganese superoxide dismutase (SOD) 

29 protein and significant decrements of extra-cellular SOD protein.

30 Various immunological outcomes were followed in offspring after their pregnant dams 

31 (BALB/c mice) were exposed gestationally to O3 (0, 0.4, 0.8, or 1.2 ppm, GD9-18) 

32 (Sharkhuu et al., 2011). Delayed type hypersensitivity (DTH) was initiated with initial 

33 BSA injection at 6 weeks of age and then challenge 7 days later. The normal edematous 

34 response of the exposed footpad (thickness after BSA injection) was recorded as an 

35 indicator of DTH. In female offspring, normal footpad swelling with BSA injection that 

36 was seen in air exposed animals was significantly attenuated with O3 exposure (0.8 and 

37 1.2 ppm O3), implying immune suppression of O3 exposure specifically in DTH. 
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1 Humoral immunity was measured with the sheep red blood cell (SRBC) response. 

2 Animals received primary immunization with SRBC and then blood was drawn for 

3 SRBC IgM measurement. A SRBC booster was given 2 weeks later with blood collected 

4 5 days after booster for IgG measurement. Maternal O3 exposure had no effect on 

5 humoral immunity in the offspring as measured by IgG and IgM titers after SRBC 

6 primary and booster immunizations (Sharkhuu et al., 2011).

7 Toxicity assessment and allergen sensitization was also assessed in these O3 exposed 

8 offspring. At PND42, animals were euthanized for analysis of immune and inflammatory 

9 markers (immune proteins, inflammatory cells, T cell populations in the spleen). A subset 

10 of the animals was intra-nasally instilled or sensitized with ovalbumin on either PND2 

11 and 3 or PND42 and 43. All animals were challenged with OVA on PND54, 55, and 56. 

12 One day after final OVA challenge, lung function, lung inflammation and immune 

13 response were determined. Offspring of O3 exposed dams that were initially sensitized at 

14 PDN3 (early) or PND42 (late) were tested to determine the level of allergic sensitization 

15 or asthma-like inflammation after OVA challenge. Female offspring sensitized early in 

16 life developed significant eosinophilia (1.2 ppm O3) and elevated serum OVA-specific 

17 IgE (1.2 ppm O3), which is a marker of airway allergic inflammation. The females that 

18 were sensitized early also had significant decrements in BALF total cells, macrophages, 

19 and lymphocytes (1.2 ppm O3). Offspring that were sensitized later (PND42) in life did 

20 not develop the aforementioned changes in BALF, but these animals did develop modest, 

21 albeit significant neutropenia (0.8 and 1.2 ppm O3) (Sharkhuu et al., 2011).

22 BALF cytology in non-sensitized animals was followed. BALF of offspring born to dams 

23 exposed to O3 was relatively unaffected (cytokines, inflammatory cell numbers/types) as 

24 were splenic T cell subpopulations. LDH was significantly elevated in BALF of females 

25 whose mothers were exposed to 1.2 ppm during pregnancy (Sharkhuu et al., 2011). In 

26 summary, the females born to mothers exposed to O3 developed modest 

27 immunocompromise. Males were unaffected (Sharkhuu et al., 2011).

28 Overall, animal toxicological studies have reported ultrastructural changes in bronchiole 

29 development, alterations in placental and pup cytokines, and increased pup airway hyper30 reactivity related to exposure to O3 during the developmental period. Epidemiologic 

31 studies have found no association between prenatal exposure to O3 and growth and 

32 development of the respiratory system. Fetal origins of disease have received a lot of 

33 attention recently, thus additional research to further explore the inconsistencies between 

34 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 

11 to O3 during fetal and neonatal life showed limited, gender-specific changes in 

12 handedness after exposure to the intermediate dose of O3 (only seen in female mice 

13 exposed 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

14 with 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 

27 offspring outcomes (i.e., dam behavioral or lactational contributions to pup outcomes) 

28 (Boussouar 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). Similar to this and 

24 as reported in previous AQCDs, continuous exposure of adult animals to O3 induced 

25 significant increases in fear behavior and decreased aggression as measured by 

26 significantly decreased freezing behavior (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.
Draft – Do Not Cite or Quote 7-63 September 2011

7.4.9.4 Sleep Aberrations after Developmental Ozone Exposure

1 The effect of gestational O3 exposure (1 ppm O3, 12h/day, during dark period) on sleep 

2 patterns in rat offspring was followed using 24 h polysomnographic recordings at 30, 60 

3 and 90 days of age (Haro and Paz, 1993). Ozone-exposed pups manifested with inverted 

4 sleep-wake patterns or circadian rhythm phase-shift. Rat vigilance was characterized in 

5 wakefulness, slow wave sleep (SWS), and paradoxical sleep (PS) using previously 

6 characterized criteria. The O3 exposed offspring spent longer time in the wakefulness 

7 state during the light period, more time in SWS during the period of darkness, and 

8 showed significant decrements in PS. Chronic O3 inhalation significantly decreased the 

9 duration of PS during both the light and dark periods (Haro and Paz, 1993). These effects 

10 were consistent at all time periods measured (30, 60 and 90 days of age). These sleep 

11 effects reported after developmental exposures expand upon the existing literature on 

12 sleep aberrations in adult animals exposed to O3 [rodents: (Paz and Huitron-Resendiz, 

13 1996; Arito et al., 1992); and cats: (Paz and Bazan-Perkins, 1992)]. A role for inhibition 

14 of cyclooxygenase-2 and the interleukins and prostaglandins in the O3-dependent sleep 

15 changes potentially exists with evidence from a publication on indomethacin pretreatment 

16 attenuating O3-induced sleep aberrations in adult male animals (Rubio and Paz, 2003). 

7.4.10 Early Life Mortality

17 Infants may be particularly susceptible to the adverse effects of air pollution. Within the 

18 first year of life, infants develop rapidly; therefore their susceptibility may change within 

19 weeks or months. During the neonatal and post-neonatal periods, the developing lung is 

20 highly susceptible to environmental toxicants. The lung is not well developed at birth, 

21 with 80% of alveoli being formed postnatally. An important question regarding the 

22 association between O3 and infant mortality is the critical window of exposure during 

23 development for which infants are susceptible. Several age intervals have been explored: 

24 neonatal (<1 month); postneonatal (1 month to 1 year); and an overall interval for infants 

25 that includes both the neonatal and postneonatal periods (<1 year). Within these various 

26 age categories, multiple causes of deaths have been investigated, particularly total deaths 

27 and respiratory-related deaths. The studies reflect a variety of study designs, exposure 

28 periods, regions, and adjustment for confounders. As discussed below, a handful of 

29 studies have examined the effect of ambient air pollution on neonatal and postneonatal 

30 mortality, with the former the least studied. These studies varied somewhat with regard to 

31 the outcomes and exposure periods examined and study designs employed.

32 A major issue in studying environmental exposures and infant mortality is selecting the 

33 relevant exposure period, since the biological mechanisms leading to death and the 
Draft – Do Not Cite or Quote 7-64 September 2011

1 critical periods of exposure are poorly understood. Both short-term (days to weeks) and 

2 long-term (months to years) exposure studies are included in this section and are 

3 characterized accordingly in the text and tables. All studies of infant mortality are 

4 included in the Reproductive and Developmental Effects section, as opposed to the 

5 sections devoted to all- and cause-specific mortality, because infant development 

6 processes, much like fetal development processes, may be particularly susceptible to O3-

7 induced health effects. Exposures proximate to the death may be most relevant if 

8 exposure causes an acute effect. However, exposure occurring in early life might affect 

9 critical growth and development, with results observable later in the first year of life, or 

10 cumulative exposure during the first year of life may be the most important determinant. 

11 The studies reviewed below have dealt with this issue in different ways. Many have 

12 considered several exposure metrics based on different periods of exposure. 

7.4.10.1 Stillbirth 

13 Pereira et al. (1998) investigated the association among daily counts of intrauterine 

14 mortality (over 28 weeks of gestation) and air pollutant concentrations in Sao Paulo, 

15 Brazil from 1991 through 1992. The association was strong for NO2, but lesser for SO2

16 and CO. These associations exhibited a short lag time, less than 5 days. No significant 

17 association was detected between short-term O3 exposure and intrauterine mortality.

7.4.10.2 Infant Mortality, Less than 1 Year 

18 Ritz et al. (2006) linked birth and death certificates for infants who died between 1989 

19 and 2000 to evaluate the influence of outdoor air pollution on infant death in the South 

20 Coast Air Basin of California. The authors examined short- and long-term exposure 

21 periods 2 weeks, 1 month, 2 months, and 6 months before each case subject’s death and 

22 reported no association between ambient levels of O3 and infant mortality. Similarly, 

23 Diaz et al. (2004) analyzed the effects of extreme temperatures and short-term exposure 

24 to air pollutants on daily mortality in children less than 1 year of age in Madrid, Spain, 

25 from 1986 to 1997 and observed no statistically significant association between mortality 

26 and O3 concentrations. Hajat et al. (2007) analyzed time-series data of daily infant 

27 mortality counts in 10 major cities in the UK to quantify any associations with short-term 

28 changes in air pollution. When the results from the 10 cities were combined there was no 

29 relationship between O3 and infant mortality, even after restricting the analysis to just the 

30 summer months. 
Draft – Do Not Cite or Quote 7-65 September 2011

1 Conversely, a time-series study of infant mortality conducted in the southwestern part of 

2 Mexico City in the years 1993-1995 found that infant mortality was associated with 

3 short-term exposure to NO2 and O3 3-5 days before death, but not as consistently as with 

4 PM. A 10-ppb increase in 24-h avg O3 was associated with a 2.78% increase (95% CI: 

5 0.29, 5.26%) in infant mortality (lag 3) (Loomis et al., 1999). This increase was 

6 attenuated, although still positive when evaluated in a two-pollutant model with PM2.5. 

7 One-hour max concentrations of O3 exceeded prevailing Mexican and international 

8 standards nearly every day. 

7.4.10.3 Neonatal Mortality, Less than 1 Month

9 Several studies have evaluated ambient O3 concentrations and neonatal mortality and 

10 observed no association. Ritz et al. (2006) linked birth and death certificates for infants 

11 who died between 1989 and 2000 to evaluate the influence of outdoor air pollution on 

12 infant death in the South Coast Air Basin of California. The authors examined short- and 

13 long-term exposure periods 2 weeks, 1 month, 2 months, and 6 months before each case 

14 subject’s death and reported no association between ambient levels of O3 and neonatal 

15 mortality. Hajat et al. (2007) analyzed time-series data of daily infant mortality counts in 

16 10 major cities in the UK to quantify any associations with short-term changes in air 

17 pollution. When the results from the 10 cities were combined there was no relationship 

18 between O3 and neonatal mortality, even after restricting the analysis to just the summer 

19 months. Lin et al. (2004a) assessed the impact of short-term changes in air pollutants on 

20 the number of daily neonatal deaths in Sao Paulo, Brazil. The authors observed no 

21 association between ambient levels of O3 and neonatal mortality. 

7.4.10.4 Postneonatal Mortality, 1 Month to 1 Year

22 A number of studies focused on the postneonatal period when examining the effects of 

23 O3 on infant mortality. Ritz et al. (2006) linked birth and death certificates for infants 

24 who died between 1989 and 2000 to evaluate the influence of outdoor air pollution on 

25 infant death in the South Coast Air Basin of California. The authors examined short- and 

26 long-term exposure periods 2 weeks, 1 month, 2 months, and 6 months before each case 

27 subject’s death and reported no association between ambient levels of O3 and 

28 postneonatal mortality. Woodruff et al. (2008) evaluated the county-level relationship 

29 between cause-specific postneonatal infant mortality and long-term early-life exposure 

30 (first 2 months of life) to air pollutants across the U.S. Similarly, they found no 

31 association between O3 exposure and deaths from respiratory causes. In the U.K., Hajat 

32 et al. (2007) analyzed time-series data of daily infant mortality counts in 10 major cities 
Draft – Do Not Cite or Quote 7-66 September 2011

1 to quantify any associations with short-term changes in air pollution. When the results 

2 from the 10 cities were combined there was no relationship between O3 and postneonatal 

3 mortality, even after restricting the analysis to just the summer months. In Ciudad Juarez, 

4 Mexico, Romieu et al. (2004b) examined the daily number of deaths between 1997 and 

5 2001, estimating the modifying effect of SES on the risk of postneonatal mortality. 

6 Ambient O3 concentrations were not related to infant mortality overall, or in any of the 

7 SES groups. In a follow-up study, Carbajal-Arroyo (2011) evaluated the relationship of 

1-h daily max O3 8 levels with postneonatal infant mortality in the Mexico City 

9 Metropolitan Area between 1997 and 2005. Generally, short-term exposure to O3 was not 

10 significantly related to infant mortality. However, upon estimating the modifying effect 

11 of SES on the risk of postneonatal mortality, the authors found that O3 was statistically 

12 significantly related to respiratory mortality among those with low SES. In a separate 

13 analysis, the effect of PM10 was evaluated with O3 level quartiles. PM10 alone was related 

14 to a significant increase in all-cause mortality. The magnitude of this effect remained the 

15 same when only the days when O3 was in the lowest quartile were included in the 

16 analyses. However, when only the days when O3 was in the highest quartile were 

17 included in the analyses, the magnitude of the PM10 effect increased dramatically 

18 (OR=1.06 [95% CI: 0.909, 1.241] for PM10 on days with O3 in lowest quartile; OR=1.26 

19 [95% CI: 1.08, 1.47] for PM10 on days with O3 in the highest quartile. These results 

20 suggest that while O3 alone may not have an effect on infant mortality, it may serve to 

21 potentiate the observed effect of PM10 on infant mortality.

22 Tsai et al. (2006) used a case-crossover analysis to examine the relationship between 

23 short-term exposure to air pollution and postneonatal mortality in Kaohsiung, Taiwan 

24 during the period 1994-2000. The risk of postneonatal deaths was 1.023 (95% CI: 0.564, 

25 1.858) per 10-ppb increase in 24-h avg O3. The confidence interval for this effect 

26 estimate is very wide, likely due to the small number of infants that died each day, 

27 making it difficult to interpret this result. Several other studies conducted in Asia did not 

28 find any association between O3 concentrations and infant mortality in the postneonatal 

29 period. Ha et al. (2003) conducted a daily time-series study in Seoul, Korea to evaluate 

30 the effect of short-term changes in ambient 8-h O3 concentrations on postneonatal 

31 mortality. Son et al. (2008) examined the relationship between air pollution and 

32 postneonatal mortality from all causes among firstborn infants in Seoul, Korea during 

33 1999-2003. Yang et al. (2006) used a case-crossover analysis to examine the relationship 

34 between air pollution exposure and postneonatal mortality in Taipei, Taiwan for the 

35 period 1994-2000. The authors observed no associations between ambient levels of O3

36 and postneonatal mortality. 
Draft – Do Not Cite or Quote 7-67 September 2011

7.4.10.5 Sudden Infant Death Syndrome

1 The strongest evidence for an association between ambient O3 concentrations and SIDS 

2 comes from a study that evaluated the county-level relationship between SIDS and long3 term early-life exposure (first 2 months of life) to air pollutants across the U.S.(Woodruff 

4 et al., 2008). The authors observed a 1.20 (95% CI: 1.09, 1.32) odds ratio for a 10-ppb 

5 increase in O3 and deaths from SIDS. There was a monotonic increase in odds of SIDS 

6 for each quartile of O3 exposure compared with the lowest quartile (highest quartile 

7 OR=1.51; [95% CI: 1.17, 1.96]). In a multi-pollutant model including PM10 or PM2.5, CO 

8 and SO2, the OR for SIDS and O3 was not substantially lower than that found in the 

9 single-pollutant model. When examined by season, the relationship between SIDS deaths 

10 and O3 was generally consistent across seasons with a slight increase for those babies 

11 born in the summer. When stratified by birth weight, the OR for LBW babies was 1.27 

12 (95% CI: 0.95, 1.69) per 10-ppb increase in O3 and the OR for normal weight babies was 

13 1.16 (95% CI: 1.01, 1.32) per 10-ppb increase in O3. 

14 Conversely, two additional studies reported no association between ambient levels of O3 

15 and SIDS. Ritz et al. (2006) linked birth and death certificates for infants who died 

16 between 1989 and 2000 to evaluate the influence of outdoor air pollution on infant death 

17 in the South Coast Air Basin of California. The authors examined short- and long-term 

18 exposure periods 2 weeks, 1 month, 2 months, and 6 months before each case subject’s 

19 death and reported no association between ambient levels of O3 and SIDS. Dales et al. 

20 (2004) used time-series analyses to compare the daily mortality rates for SIDS and short21 term air pollution concentrations in 12 Canadian cities during the period of 1984-1999. 

22 Increased daily rates of SIDS were associated with previous day increases in the levels of 

23 SO2, NO2, and CO, but not O3 or PM2.5. 

24 Table 7-9 provides a brief overview of the epidemiologic studies of infant mortality. 

25 These studies have focused on short-term exposure windows (e.g., 1-3 days) and long26 term exposure windows (e.g., up to 6 months). Collectively, they provide no evidence for 

27 an association between ambient O3 concentrations and infant mortality. 
Draft – Do Not Cite or Quote 7-68 September 2011

Table 7-9 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 wk 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. (2004b) 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)

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, US 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
Draft – Do Not Cite or Quote 7-69 September 2011

Table 7-10 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. 

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, 10 or 18)

3.0 1h 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 coadministered 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. 
Draft – Do Not Cite or Quote 7-70 September 2011

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 

3 exception of birth defects. The current review included an expanded body of evidence on 

4 the 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, 

8 alterations in placental and pup cytokines, and increased pup airway hyper-reactivity. 

9 Also, there is limited toxicological evidence for an effect of prenatal and early life 

10 exposure on central nervous system effects, including laterality, brain morphology, 

11 neurobehavioral abnormalities, and sleep aberration. Recent epidemiologic studies have 

12 begun to explore the effects of O3 on sperm quality, and provide limited evidence for 

13 decrements in sperm concentration, while there is limited toxicological evidence for 

14 testicular degeneration associated with O3. 

15 While the collective evidence for many of the birth outcomes examined is generally 

16 inconsistent (including birth defects), there are several well-designed, well-conducted 

17 studies that indicate an association between O3 and adverse outcomes. For example, as 

18 part of the southern California Children’s Health Study, Salam et al. (2005) observed a 

19 concentration-response relationship of decreasing birth weight with increasing O3

20 concentrations averaged over the entire pregnancy that was clearest above the 30-ppb

21 level (see Figure 7-4). Simiarly,Hansen et al. (2008) utilized fetal ultrasonic 

22 measurements and found a change in ultrasound measurements associated with O3 during 

23 days 31-60 of gestation indicated that increasing O3 concentration decreased an 

24 ultrasound measurement for women living within 2 km of the monitoring site.

25 There is no evidence that prenatal or early life O3 concentrations are associated with 

26 infant mortality. Collectively, there is limited though positive toxicological evidence for 

27 O3-induced developmental effects, including effects on pulmonary structure and function 

28 and central nervous system effects. Limited epidemiologic evidence for an effect on 

29 prenatal O3 exposure on respiratory development provides coherence with the effects 

30 observed in toxicological studies. There is also limited epidemiologic evidence for an 

31 association with O3 concentration and decreased sperm concentration. A recent 

32 toxicological study provides limited evidence for a possible biological mechanism 

33 (histopathology showing impaired spermatogenesis) for such an association. 

34 Additionally, though the evidence for an association between O3 concentrations and 

35 adverse birth outcomes is generally inconsistent, there are several influential studies that 

36 indicate an association with reduced birth weight and restricted fetal growth. Taking into 
Draft – Do Not Cite or Quote 7-71 September 2011

1 consideration the positive evidence for developmental and reproductive outcomes from 

2 toxicological and epidemiological studies, and the few influential birth outcome studies, 

3 the evidence is suggestive of a causal relationship between long-term exposures to O3

4 and reproductive and developmental effects.

7.5 Central Nervous System Effects 

7.5.1 Effects on the Brain and Behavior

5 The 2006 O3 AQCD included toxicological evidence that acute exposures to O3 are 

6 associated with alterations in neurotransmitters, motor activity, short and long term 

7 memory, and sleep patterns. Additionally, histological signs of neurodegeneration have 

8 been observed. Reports of headache, dizziness, and irritation of the nose with O3

9 exposure are common complaints in humans, and some behavioral changes in animals 

10 may be related to these symptoms rather than indicative of neurotoxicity. Research in the 

11 area of O3-induced neurotoxicity has notably increased over the past few years, and new 

12 studies examining the effects of long-term exposure have demonstrated progressive 

13 damage in various regions of the brains of rodents in conjunction with altered behavior. 

14 Evidence from epidemiologic studies has been more limited. A recently published 

15 epidemiologic study examined the association between O3 exposure and neurobehavioral 

16 effects. Chen et al. (2009) utilized data from the NHANES III cohort to study the 

17 relationship between O3 levels (mean annual O3 concentration 26.5 ppb) and 

18 neurobehavioral effects among adults aged 20-59 years. The authors observed an 

19 association between annual exposure to O3 and tests measuring coding ability (symbol20 digit substitution test) and attention/short-term memory (serial-digit learning test). Each 

21 10-ppb increase in annual O3 levels corresponded to an aging-related cognitive 

22 performance decline of 3.5 yr for coding ability and 5.3 years for attention/short-term 

23 memory. These associations persisted in both crude and adjusted models. There was no 

24 association between O3 levels and reaction time tests. The authors concluded that overall, 

25 there is an association between long-term O3 exposure and reduced performance on 

26 neurobehavioral tests.

27 A number of new toxicological studies demonstrate various perturbations in neurologic 

28 function or histology with long-term exposure to O3, including changes similar to those 

29 observed in neurodegenerative disorders such as Parkinson’s and Alzheimer’s disease 

30 pathologies in relevant regions of the brain (Table 7-11). The central nervous system is 

31 very sensitive to oxidative stress, due in part to its high content of polyunsaturated fatty 

32 acids, high rate of oxygen consumption, and low antioxidant enzyme capacity. Oxidative 
Draft – Do Not Cite or Quote 7-72 September 2011

1 stress has been identified as one of the pathophysiological mechanisms underlying 

2 neurodegenerative disease (Simonian and Coyle, 1996), and it is believed to play a role in 

3 altering hippocampal function, which causes cognitive deficits with aging (Vanguilder 

4 and Freeman, 2011). A particularly common finding in studies of O3-exposed rats is lipid 

5 peroxidation in the brain, especially in the hippocampus, which is important for higher 

6 cognitive function including contextual memory acquisition. Performance in passive 

7 avoidance learning tests is impaired when the hippocampus is injured. For example, in a 

8 subchronic study, exposure of rats to 0.25 ppm O3 (4 h/day) for 15-90 days caused a 

9 complex array of responses, including a time-dependent increase in lipid peroxidation 

10 products and immunohistochemical changes in the hippocampus that were correlated 

11 with decrements in passive avoidance behavioral tests (Rivas-Arancibia et al., 2010). 

12 Changes included increased numbers of activated microglia, a sign of inflammation, and 

13 progressive neurodegeneration. Notably, continued exposure tends to bring about 

14 progressive, cumulative damage, as shown by this study (Rivas-Arancibia et al., 2010)

15 and others (Santiago-López et al., 2010; Guevara-Guzmán et al., 2009; Angoa-Pérez et 

16 al., 2006). The effects of O3 on passive avoidance test performance were particularly 

17 evident at 90 days for both short- and long-term memory. The greatest extent of cell loss 

18 was also observed at this time point, whereas lipid peroxidation did not increase much 

19 beyond 60 days of exposure.

20 The substantia nigra is another region of the brain affected by O3, and seems particularly 

21 sensitive to oxidative stress because the metabolism of dopamine, central to its function, 

22 is an oxidative process perturbed by redox imbalance. Oxidative stress has been 

23 implicated in the premature death of substantia nigra dopamine neurons in Parkinson’s 

24 disease. Progressive damage has been found in the substantia nigra of male rats after 15, 

25 30, and 60 days of exposure to 0.25 ppm O3 for 4 h/day. Santiago-López and colleagues 

26 (2010) observed a reduction dopaminergic neurons within the substantia nigra over time, 

27 with a complete loss of normal morphology in the remaining cells and virtually no

28 dopamine immunoreactivity at 60 days. This was accompanied by an increase in p53 

29 levels and nuclear translocation, a process associated with programmed cell death. 

30 Similarly, Angoa-Pérez et al. (2006) have shown progressive lipoperoxidation in the 

31 substantia nigra and a decrease in nigral neurons in ovariectomized female rats exposed 

32 to 0.25 ppm O3, 4h/day, for 7 - 60 days. Lipid peroxidation effectively doubled between 

33 the 30 and 60 day time points. Total nigral cell number was also diminished to the 

34 greatest extent at 60 days, and cell loss was particularly evident in the tyrosine 

35 hydroxylase positive cell population (90%), indicating a selective loss of dopamine 

36 neurons or a loss of dopamine pathway functionality.

37 The olfactory bulb also undergoes oxidative damage in O3-exposed animals, in some 

38 cases altering olfactory-dependent behavior. Lipid peroxidation was observed in the 
Draft – Do Not Cite or Quote 7-73 September 2011

1 olfactory bulbs of ovariectomized female rats exposed to 0.25 ppm O3 (4 h/day) for 30 or 

2 60 days (Guevara-Guzmán et al., 2009). O3 also induced decrements in a selective 

3 olfactory recognition memory test, which were significantly greater at 60 days compared 

4 to 30 days, and the authors note that early deficits in odor perception and memory are 

5 components of human neurodegenerative diseases. The decrements in olfactory memory 

6 did not appear to be due to damaged olfactory perception based on other tests early on, 

7 but by 60 days deficits in olfactory perception had emerged.

8 Memory deficits and associated morphological changes can be attenuated by 

9 administration of α-tocopherol (Guerrero et al., 1999), taurine (Rivas-Arancibia et al., 

10 2000), and estradiol (Guevara-Guzmán et al., 2009; Angoa-Pérez et al., 2006), all of 

11 which have antioxidant properties. In the study by Angoa-Pérez et al. (2006) described 

12 above, estradiol seemed particularly effective at protecting against lipid peroxidation and 

13 nigral cell loss at 60 days compared to shorter exposure durations. The same was true for 

14 amelioration of decrements in olfactory recognition memory (Guevara-Guzmán et al., 

15 2009), although protection against lipid peroxidation was similar for the 30 and 60 day 

16 exposures. 

Table 7-11 Central nervous system effects of long-term O3 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/wk

Long-term estradiol treatment protected against 

O3-induced oxidative damage to nigral 

dopamine neurons, lipid peroxidation, and loss 

of tyrosine hydrolase-immunopositive cells. 

Guevara-Guzmán 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 βhydroxlyase 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 memory-retention 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 nose-sniffing. These behavioral 

changes were accompanied by a significant 

increase in BDNF in the striatum and a 

decrease of NGF in the hippocampus.
Draft – Do Not Cite or Quote 7-74 September 2011

1 CNS effects have also been demonstrated in adult mice whose only exposure to O3

2 occurred while in utero, a period particularly critical for brain development. Santucci et 

3 al. (2006) investigated behavioral effects and gene expression after in utero exposure of 

4 mice to 0.3 or 0.6 ppm O3. Exposure began 30 days prior to mating and continued 

5 throughout gestation. Testing of adult animals demonstrated increased 

6 defensive/submissive behavior and reduced social investigation were observed in both the 

7 0.3 and 0.6 ppm O3 groups. Changes in gene expression of brain-derived neurotrophic 

8 factor (BDNF, increased in striatum) and nerve growth factor (NGF, decreased in 

9 hippocampus) accompanied these behavioral changes. BDNF and NGF are involved in 

10 neuronal organization and the growth, maintenance, and survival of neurons during early 

11 development and in adulthood. This study and two others using short-term exposures 

12 demonstrate that CNS effects can occur as a result of in utero exposure to O3, and 

13 although the mode of action of these effects is not known, it has been suggested that 

14 circulating lipid peroxidation products may play a role (Boussouar et al., 2009). 

15 Importantly, these CNS effects occurred in rodent models after in utero only exposure to 

16 (semi-) relevant concentrations of O3.

7.5.2 Summary and Causal Determination 

17 The 2006 O3 AQCD included toxicological evidence that acute exposures to O3 are 

18 associated with alterations in neurotransmitters, motor activity, short and long term 

19 memory, and sleep patterns. Additionally, histological signs of neurodegeneration have 

20 been observed. However, evidence regarding chronic exposure and neurobehavioral 

21 effects was not available. Recent research in the area of O3-induced neurotoxicity has 

22 included several long-term exposure studies. Notably, the first epidemiologic study to 

23 examine the relationship between O3 exposure and neurobehavioral effects observed an 

24 association between annual O3 levels and an aging-related cognitive performance decline 

25 in tests measuring coding ability and attention/short-term memory. This observation is 

26 supported by studies in rodents which demonstrate progressive oxidative stress and 

27 damage in the brain and associated decrements in behavioral tests, including those 

28 measuring memory, after subchronic exposure to 0.25 ppm O3. Additionally, 

29 neurobehavioral changes are evident in animals whose only exposure to O3 occurred in 

30 utero. Collectively, the limited epidemiologic and toxicological evidence is coherent and 

31 suggestive of a causal relationship between O3 exposure and CNS effects.
Draft – Do Not Cite or Quote 7-75 September 2011

7.6 Carcinogenic and Genotoxic Potential of Ozone

7.6.1 Introduction

1 The radiomimetic and clastogenic qualities of O3, combined with its ability to stimulate 

2 proliferation of cells in the respiratory tract, have suggested that O3 could act as a 

3 carcinogen. However, toxicological studies of tumorigenesis in the rodent lung have 

4 yielded mixed and often confusing results, and the epidemiologic evidence is equally 

5 conflicted. The 2006 O3 AQCD concluded that, “the weight of evidence from recent 

6 animal toxicological studies and a very limited number of epidemiologic studies do not 

support ambient O3 as a pulmonary carcinogen” 2 7 (U.S. EPA, 2006b). 

8 Multiple epidemiologic studies reported in the 2006 O3 AQCD examined the association 

9 between O3 exposure and cancer. The largest of these studies, by Pope et al. (2002), 

10 included 500,000 adults from the American Cancer Society’s (ACS) Cancer Prevention II 

11 study. In this study, no association was observed between O3 and lung cancer mortality. 

12 The Adventist Health Study of Smog (AHSMOG) also examined the association between 

13 O3 and lung cancer mortality (Abbey et al., 1999). There was a positive association 

14 between O3 levels and lung cancer mortality among men. No association was reported for 

15 women. Another study using the AHSMOG cohort assessed the risk of incident lung 

16 cancer (Beeson et al., 1998). Among males, an association with incidence of lung cancer 

17 was observed with increasing O3 concentrations. When stratified by smoking status, the 

18 association persisted among never smokers but was null for former smokers. No 

19 association was detected for females. The Six Cities Study examined various air 

20 pollutants and mortality but did not specifically explore the association between O3

21 concentrations and lung cancer mortality due to low variability in O3 levels across the 

22 cities (Dockery et al., 1993). An ecologic study performed in Sao Paulo City, Brazil 

23 examined the correlations between O3 levels in four of the city districts and incident 

24 cancer of the larynx and lung reported in 1997 (Pereira et al., 2005). A correlation 

25 between the average number of days O3 levels exceeded air quality standards from 1981 

26 to 1990 and cancer incidence was present for larynx cancer but not for lung cancer. 

Early toxicological research demonstrated lung adenoma3 27 acceleration in mice with daily 

28 exposure to 1 ppm over 15 months (Stokinger, 1962). Later work demonstrated a 

29 significant increase in lung tumor numbers in one strain of mouse (A/J) but not another 

 2 The toxicological evidence is presented in detail in Table 6-18 on p. 6-116 of the 1996 O3 AQCD and Table AX5-13 on p. AX5-43 

of the 2006 O3 AQCD.

3 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.”
Draft – Do Not Cite or Quote 7-76 September 2011

1 after exposure to 0.3-0.8 ppm O3 (Last et al., 1987; Hassett et al., 1985). The A/J mouse 

2 strain is known to have a high incidence of spontaneous adenomas, and further studies 

3 using this strain found a statistically significant increase in lung tumor incidence after a 

4 9-month exposure to 0.5 ppm and incidence and multiplicity after a 5 month exposure to 

5 0.12 ppm with a 4-month recovery period (Witschi et al., 1999). However, these findings 

6 were discounted by the study authors due to the lack of a clear dose response, and results 

7 from the Hassett et al. 1985 and Last et al. 1987 studies were retrospectively deemed 

8 spurious based on what appeared to be unusually low spontaneous tumor incidences in 

9 the control groups (Witschi, 1991). A study of carcinogenicity of O3 by the National 

10 Toxicology Program (NTP, 1994) reported increased incidences of alveolar/bronchiolar 

11 adenoma or carcinoma (combined) in female B6C3F1 mice exposed over 2 years to 

12 1.0 ppm O3, but not 0.12 or .5 ppm. No effect was detected in male mice. For a lifetime 

13 exposure to 0.5 or 1.0 ppm O3, an increase in the number of female mice with adenomas 

14 (but not carcinomas or total neoplasms) was found. The number of total neoplasms was 

15 also unaffected in male mice, but there was a marginally increased incidence of 

16 carcinoma in males exposed to 0.5 and 1.0 ppm. Thus there was equivocal evidence of

17 carcinogenic activity in male mice and some evidence of carcinogenic activity of O3 in 

18 females. Some semblance of a dose-response relationship was also evident in this study. 

19 Experimental details of the NTP study are available in Table 6-19 on p. 6-121 of the 1996 

20 O3 AQCD.

21 In Fischer-344/N rats (50 of each sex per group), neither a 2-year nor lifetime exposure to 

22 O3 ranging from 0.12 to 1.0 ppm was found to be carcinogenic (Boorman et al., 1994). 

23 However, a marginally significant carcinogenic effect of 0.2 ppm O3 was reported in a 

24 study of male Sprague-Dawley rats exposed for 6 months (n = 50) (Monchaux et al., 

1996). These two studies also examined co-carcinogenicity of O3 with NNK4 25 (Boorman 

26 et al., 1994) or a relatively high dose of radon (Monchaux et al., 1996), finding no 

27 enhancement of NNK related tumors and a slight non-significant increase in tumor 

28 incidence after combined exposure with radon, respectively. Another study exploring co29 carcinogenicity was conducted in hamsters. Not only was there no enhancement of 

30 chemically induced tumors in the peripheral lung or nasal cavity, but results suggested 

31 that O3 could potentially delay or inhibit tumor development (Witschi et al., 1993). Thus 

32 there is no concrete evidence that O3 can act as a co-carcinogen. 

33 Immune surveillance is an important defense against cancer, and it should be noted that 

34 natural killer (NK) cells, which destroy tumor cells in the lung, appear to be inhibited by 

35 higher doses of O3 and either unaffected or stimulated at lower doses (Section 6.2.5.4, 

 4 4-(N-nitrosomethylamino)-1-(3-pyridyl)-1-butanone
Draft – Do Not Cite or Quote 7-77 September 2011

1 Infection and Adaptive Immunity). This aspect of tumorigenesis adds yet another layer of 

2 complexity which may be reflected by conflicting results across studies.

3 The following sections will examine epidemiologic studies of cancer incidence and 

4 mortality and toxicological studies that have been published since the 2006 O3 AQCD. 

5 An epidemiologic study has been published with cancer as the outcome; most 

6 epidemiologic studies examine markers of exposure or susceptibility. 

7.6.2 Lung Cancer Incidence and Mortality

7 A recent re-analysis of the full ACS CPSII cohort by the Health Effects Institute is the 

8 only epidemiologic study that has explored the association between O3 and cancer 

9 mortality since the last O3 AQCD. Krewski et al. (2009) conducted an extended follow10 up of the cohort (1982-2000). Mean O3 levels [obtained from the Aerometric Information 

11 Retrieval System (AIRS) for 1980] were 22.91 ppb for the full year and 30.15 ppb for the 

12 summer months (April-September). No association was reported between lung cancer 

mortality and O3 (HR=1.00 [95% CI: 0.96-1.04] per 10 ppb O3) 13 . Additionally, no 

14 association was observed when O3 was restricted to the summer months. There was also 

15 no association present in a sub-analysis of the cohort examining the relationship between 

16 O3 and lung cancer mortality in the Los Angeles area.

17 Since the 2006 O3 AQCD, two toxicological studies have examined potential 

18 carcinogenicity of O3 (Kim and Cho, 2009a, b). Looking across both studies, which used 

19 the same mouse strain as the National Toxicology Program study described above (NTP, 

20 1994), 0.5 ppm O3 alone or in conjunction with chemical tumor inducers did not enhance 

21 lung tumor incidence in males or females. However, a 10% incidence of oviductal 

22 carcinoma was observed in mice exposed to 0.5 ppm O3 for 16 weeks. The implications 

23 of this observation are unclear, particularly in light of the lack of statistical information 

24 reported. Additionally, there is no mention of oviductal carcinoma after 32 weeks of 

25 exposure, and no oviductal carcinoma was observed after one year of exposure. The NTP 

26 study did not report any increase in tumors at extrapulmonary sites.

7.6.3 DNA Damage

27 The potential for genotoxic effects relating to O3 exposure was predicted from the 

28 radiomimetic properties of O3. The decomposition of O3 in water produces OH and HO2

29 radicals, the same species that are generally considered to be the biologically active 

30 products of ionizing radiation. Ozone has been observed to cause degradation of DNA in 

31 a number of different models and bacterial strains. The toxic effects of O3 have been 
Draft – Do Not Cite or Quote 7-78 September 2011

1 generally assumed to be confined to the tissues directly in contact with the gas, such as 

2 the respiratory epithelium. Due to the highly reactive nature of O3, little systemic 

3 absorption is predicted. Zelac et al. (1971a, b), however, reported a significant increase in 

4 chromosome aberrations in peripheral blood lymphocytes from Chinese hamsters 

5 exposed to 0.2 ppm for 5 hours. Other in vivo exposure studies found increased DNA 

6 strand breaks in respiratory cells from guinea pigs (Ferng et al., 1997) and mice 

7 (Bornholdt et al., 2002) but only with exposure to higher doses of O3 (1 ppm for 72 hours 

8 and 1 or 2 ppm for 90 minutes, respectively). In other studies there were no observations 

9 of chromosomal aberrations in germ cells, but mutagenic effects have been seen in 

10 offspring of mice exposed to 0.2 ppm during gestation (blepharophimosis or dysplasia of 

11 the eyelids). The overall evidence for mutagenic activity from in vitro studies is positive, 

12 and in the National Toxicology Program report described above, O3 was found to be 

13 mutagenic in Salmonella, with and without S9 metabolic activation. No new 

14 toxicological studies of DNA damage have become available since the 2006 O3 AQCD.

15 A number of epidemiologic studies looked at the association between O3 and DNA and 

16 cellular level damages. These changes may be relevant to mechanisms leading to cancers 

17 development and serve as early indicators of elevated risk of mutagenicity. 

18 Two studies performed in California examined cytogenetic damage in relation to O3

19 exposures. Huen et al. (2006) examined cytogenetic damage among African American 

20 children and their mothers in Oakland, CA. Increased O3 (mean monthly 8-h O3

21 concentrations ranged from about 30 ppb in April to 14 ppb in November) was associated 

22 with increased cytogenetic damage (micronuclei frequency among lymphocytes and 

23 buccal cells) even after adjustment for household/personal smoking status and distance24 weighted traffic density. Chen et al. (2006a) recruited college students at the University 

25 or California, Berkeley who reported never smoking and compared their levels of 

26 cytogenetic damage (micronuclei frequency from buccal cells) in the spring and fall. 

27 Cytogenetic damage was greater in the fall, which the authors attributed to the increase in 

28 O3 over the summer. However, O3 levels over 2, 7, 10, 14, or 30 days (concentrations not 

29 given) before collection of buccal cells did not correlate with cytogenetic damage. 

30 Estimated lifetime O3 exposure was also not correlated with cytogenetic damage. 

31 Additionally, the authors exposed a subset of the students (n=15) to 200 ppb O3 for 

32 4 hours while the students exercised intermittently. Ozone was found to be associated 

33 with an increase in cytogenetic damage in degenerated cells but not in normal cells 9-

34 10 days after exposure. Increased cytogenetic damage was also noted in peripheral blood 

35 lymphocytes collected 18 hours after exposure. 

36 A study performed in Mexico recruited 55 male workers working indoors (n=27) or 

37 outdoors (n=28) in Mexico City or Puebla, Mexico in order to study the relationship 
Draft – Do Not Cite or Quote 7-79 September 2011

1 between O3 and DNA damage (detected from peripheral blood samples using the Comet 

2 assay) (Tovalin et al., 2006). The median estimated daily O3 concentrations were 

3 estimated to be 28.5 ppb for outdoor workers and 5.1 ppb for indoor workers in 

4 Mexico City and 36.1 ppb for outdoor workers and 19.5 ppb for indoor workers in 

5 Puebla. Overall, a positive correlation between O3 levels and DNA damage was 

6 observed. However, when examining the relationship by city and workplace, only DNA 

7 damage in outdoor workers in Mexico City remained correlated with O3 levels. 

8 Three studies examining the relationship between O3 and DNA-level damage have been 

9 performed in Europe. The largest of these was the GenAir case-control study, which was 

10 nested within the European Prospective Investigation into Cancer and Nutrition (EPIC) 

11 study, and included individuals recruited between 1993 and 1998 from ten European 

12 countries. Only non-smokers (must not have smoked for at least 10 years prior to 

13 enrollment) were enrolled in the study. The researchers examined DNA adduct levels 

14 (DNA bonded to cancer-causing chemicals) and their relationship with O3 concentrations 

15 (concentrations not given) (Peluso M Hainaut et al., 2005). A positive association was 

16 seen between DNA adduct levels and O3 concentrations from 1990-1994 but not O3

17 levels from 1995-1999. In adjusted analyses with DNA adduct levels dichotomized as 

18 high and low (detectable versus non-detectable), the OR was 1.97 (95% CI: 1.08, 3.58) 

19 when comparing the upper tertile of O3 concentration to the lower two tertiles. Two other 

20 European studies were conducted in Florence, Italy. The most recent of these enrolled 

21 individuals from the EPIC study into a separate study between March and September of 

22 1999 (Palli et al., 2009). The purpose of the study was to examine oxidative DNA 

23 damage (determined by Comet assay using blood lymphocytes) in association with 

24 varying periods of O3 exposure. The researchers observed that longer periods of high O3

25 exposure (concentrations not given) were more strongly correlated with oxidative DNA 

26 damage than shorter exposures (i.e., the rho [p-value] was 0.26 [0.03] for 0-10 days and 

27 0.35 [0.002] for 0-90 days). This correlation was stronger among men compared to 

28 women. The correlations for all time periods had p-values <0.05 for ex- and never29 smokers. For current smokers, the correlation was only observed among time periods ≤ 

30 25 days. When adjusted for age, gender, smoking history, traffic pollution exposure, 

31 period of blood draw, and area of residence, the association between O3 levels and 

32 oxidative DNA damage was positive for O3 levels 0-60 days, 0-75 days, and 0-90 days 

33 prior to blood draw. Positive, statistically significant associations were not observed 

34 among shorter time periods. The other study performed in Florence recruited healthy 

35 volunteers who reported being non-smokers or light smokers (Giovannelli et al., 2006). 

36 The estimated O3 levels during the study ranged from approximately 4-40 ppb for 3-day 

37 avgs, 5-35 ppb for 7-day avgs, and 7.5-32.5 ppb for 30-day avgs. Ozone concentrations 

38 were correlated with DNA strand breaks (measured from blood lymphocytes) over longer 

39 exposure periods (p-value: 0.002 at 30 days, p-value: 0.04 at 7 days; p-value: 0.17 at 
Draft – Do Not Cite or Quote 7-80 September 2011

1 3 days). This association was robust to control for temperature, solar radiation, gender, 

2 and age. No association was seen between O3 concentrations and measures of oxidative 

3 DNA damage at 3, 7, or 30 days.

7.6.4 Summary and Causal Determination 

4 The 2006 O3 AQCD reported that evidence did not support ambient O3 as a pulmonary 

5 carcinogen. Since the 2006 O3 AQCD, very few epidemiologic and toxicological studies 

6 have been published that examine O3 as a carcinogen, but collectively, study results 

7 indicate that O3 may contribute to DNA damage. Overall, the evidence is inadequate to 

8 determine if a causal relationship exists between ambient O3 exposures and 

9 cancer. 

7.7 Mortality

10 A limited number of epidemiologic studies have assessed the relationship between long11 term exposure to O3 and mortality in adults. The 2006 O3 AQCD concluded that an 

12 insufficient amount of evidence existed “to suggest a causal relationship between chronic 

13 O3 exposure and increased risk for mortality in humans” (U.S. EPA, 2006b). In addition 

14 to the infant mortality studies discussed in Section 7.4.9, additional studies have been 

15 conducted among adults since the last review; an ecologic study that finds no association 

16 between mortality and O3, several reanalyses of the ACS cohort, one of which 

17 specifically points to a relationship between long-term O3 exposure and an increased risk 

18 of respiratory mortality, and a study of four cohorts of persons with potentially 

19 predisposing conditions. These studies supplement the evidence from long-term cohort 

20 studies characterized in previous reviews of O3, and are summarized here briefly.

21 In the Harvard Six Cities Study (Dockery et al., 1993), adjusted mortality rate ratios were 

22 examined in relation to long-term mean O3 concentrations in six cities: Topeka, KS; St. 

23 Louis, MO; Portage, WI; Harriman, TN; Steubenville, OH; and Watertown, MA. Mean 

24 O3 concentrations from 1977 to 1985 ranged from 19.7 ppb in Watertown to 28.0 ppb in 

25 Portage. Long-term mean O3 concentrations were not found to be associated with 

26 mortality in the six cities. However, the authors noted that “The small differences in O3

27 levels among the (six) cities limited the power of the study to detect associations between 

28 mortality and O3 levels.” In addition, while total and cardio-pulmonary mortality were 

29 considered in this study, respiratory mortality was not specifically considered.

30 In a subsequent large prospective cohort study of approximately 500,000 U.S. adults, 

31 Pope et al. (2002) examined the effects of long-term exposure to air pollutants on 
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1 mortality (American Cancer Society, Cancer Prevention Study II). All-cause, 

2 cardiopulmonary, lung cancer and other mortality risk estimates for long-term O3

3 exposure are shown in Figure 7-5. While consistently positive associations were not 

4 observed between O3 and mortality (effect estimates labeled A in Figure 7-5) , the 

5 mortality risk estimates were larger in magnitude when analyses considered more 

6 accurate exposure metrics, increasing when the entire period was considered (effect 

7 estimates labeled B in Figure 7-5) and becoming marginally significant when the 

8 exposure estimate was restricted to the summer months (July to September; effect 

9 estimates labeled C in Figure 7-5), especially when considering cardiopulmonary deaths. 

10 In contrast, consistent positive and significant effects of PM2.5 were observed for both 

11 lung cancer and cardio-pulmonary mortality.

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 O3 concentration in 

the Cancer Prevention Study II by the American Cancer Society.

12 A study by Abbey et al. (1999) examined the effects of long-term air pollution exposure, 

13 including O3, on all-cause (n = 1,575), cardiopulmonary (n = 1,029), nonmalignant 

14 respiratory (n = 410), and lung cancer (n = 30) mortality in the long-term prospective 

15 Adventist Health Study of Smog (AHSMOG) of 6,338 nonsmoking, non-Hispanic white 

16 individuals living in California. A particular strength of this study was the extensive 
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1 effort devoted to assessing long-term air pollution exposures, including interpolation to 

2 residential and work locations from monitoring sites over time and space. No associations 

3 with long-term O3 exposure were observed for all cause, cardiopulmonary, and 

4 nonmalignant respiratory mortality. In a follow-up, Chen et al. (2005) utilized data from 

5 the AHSMOG study and reported no evidence of associations between long-term O3

6 exposure (mean O3 concentration 26.2 ppb) and fatal coronary heart disease. Thus, no 

7 association of chronic O3 exposure with mortality outcomes has been detected in this 

8 study. 

9 Lipfert et al. (2003, 2000) reported positive effects on all-cause mortality for peak O3

10 exposures (95th percentile levels) in the U.S. Veterans Cohort study of approximately 

11 50,000 middle-aged men recruited with a diagnosis of hypertension. The actual analysis 

12 involved smaller subcohorts based on exposure and mortality follow-up periods. Four 

13 separate exposure periods were associated with three mortality follow-up periods. For 

14 concurrent exposure periods, peak O3 was positively associated with all-cause mortality, 

15 with a 9.4% (95% CI: 0.4, 18.4) excess risk per mean 95th percentile O3 less estimated 

16 background level (not stated). “Peak” refers, in this case, to the 95th percentile of 

17 the hourly measurements, averaged by year and county. In a further analysis, Lipfert et al. 

18 (2003) reported the strongest positive association for concurrent exposure to peak O3 for 

19 the subset of subjects with low diastolic blood pressure during the 1982 to 1988 period. 

20 Two more recent studies of this cohort focused specifically on traffic density (Lipfert et 

21 al., 2006a; 2006b). Lipfert (2006b) concluded that: “Traffic density is seen to be a 

22 significant and robust predictor of survival in this cohort, more so than ambient air 

23 quality, with the possible exception of O3,” reporting a significant O3 effect even with 

24 traffic density included in the model: RR=1.080 per 40 ppb peak O3 (95% CI: 1.019, 

25 1.146). However, in Lipfert (2006a), which considers only the EPA Speciation Trends 

26 Network (STN) sites, O3 drops to non-significant predictor of total mortality for this 

27 cohort. The authors acknowledge that: “Peak O3 has been important in analyses of this 

28 cohort for previous periods, but in the STN data set, this variable has limited range and 

29 somewhat lower values and its small coefficient of variation results in a relatively large 

30 standard error.” The restriction to subjects near STN sites likely reduced the power of this 

31 analysis, though the size of the remaining subjects considered was not reported in this 

32 paper. In addition, these various Veterans Cohort studies considered only total mortality, 

33 and did not consider mortality on a by-cause basis. 

34 An ecological study in Brisbane, Australia used a geospatial approach to analyze the 

35 association of long-term exposure to gaseous air pollution with cardio-respiratory 

36 mortality, in the period 1996-2004 (Wang et al., 2009c). A generalized estimating 

37 equations model was employed to investigate the impact of NO2, O3 and SO2, but PM 

38 was not addressed. The results indicated that long-term exposure to O3 was not 
Draft – Do Not Cite or Quote 7-83 September 2011

1 associated with cardio-respiratory mortality, but the fact that this study considered only 

2 one city, and that the range of O3 exposure across that city (23.7-35.6 ppb) was low and 

3 slight in variation in comparison to the range of other pollutants across the city, limited 

4 study power. In addition, confounding factors (e.g., smoking) could not be addressed at 

5 the individual level in this ecological study. Respiratory mortality was not evaluated 

6 separately.

7 A recent study by Zanobetti and Schwartz (In Press) examined whether year-to-year 

8 variations in 8-h mean daily O3 concentrations for the summer (May-September) around 

9 their city-specific long-term trend were associated with year-to-year variations in 

10 mortality around its long-term trend. This association was examined among Medicare 

11 participants with potentially predisposing conditions, including COPD, diabetes, CHF, 

12 and MI, defined as patients discharged alive after an emergency admission for one of 

13 these four conditions. The analyses was repeated in 105 cities using available data from 

14 1985 through 2006, and the results were combined using methods previously employed 

15 by these authors (Zanobetti et al., 2008; Zanobetti and Schwartz, 2007). This study 

16 design eliminated potential confounding by factors that vary across city, which is a 

17 common concern in most air pollution cohort studies, and also avoided both confounding 

18 by cross-sectional factors that vary by city and the short-term factors that confound daily 

19 time-series studies, but are not present in annual analyses. The average 8-h mean daily 

20 summer O3 concentrations ranged from 15.6 ppb (Honolulu, HI) to 71.4 ppb 

21 (Bakersfield, CA) for the 105 cities. The authors observed associations between yearly 

22 fluctuations in summer O3 concentrations and mortality in each of the four cohorts; the 

23 hazard ratios (per 10 ppb increment) were 1.12 (95% CI: 1.06, 1.17) for the CHF cohort, 

24 1.19 (95% CI 1.12, 1.25) for the MI cohort, 1.14 (95% CI: 1.10, 1.21) for the diabetes 

25 cohort, and 1.14 (95% CI: 1.08, 1.19) for the COPD cohort. A key advantage to this study 

26 is that fluctuations from summer to summer in O3 concentrations around long-term level 

27 and trend in a specific city are unlikely to be correlated with most other predicators of 

28 mortality risk, except for temperature, which was controlled for in the regression. Key 

29 limitations of the study were the inability to control for PM2.5, since it was not reliably 

30 measured in these cities until 1999, and the inability to separate specific causes of death 

31 (e.g., respiratory, cardiovascular), since Medicare does not provide the underlying cause 

32 of death.

33 In the most recent follow-up analyses of the ACS cohort (Jerrett et al., 2009; Smith et al., 

34 2009a), the effects of long-term exposure to O3 were evaluated alone, as well as in 

35 copollutant models with PM2.5 and components of PM2.5. Jerrett et al. (2009) utilized the 

36 ACS cohort with data from 1977 through 2000 (mean O3 concentration ranged from 33.3 

37 to 104.0 ppb) and subdivided cardiopulmonary deaths into respiratory and cardiovascular, 

38 separately, as opposed to combined into one category, as was done by Pope et al. (2002). 
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1 Increases in exposure to O3 were associated with an elevated risk of death from 

2 cardiopulmonary, cardiovascular, ischemic heart disease, and respiratory causes. 

3 Inclusion of PM2.5 concentrations measured in 1999-2000 as a copollutant attenuated the 

4 association with O3 for all end points except death from respiratory causes, for which a 

5 significant association persisted (Table 7-12). The association between increased O3

6 concentrations and increased risk of death from respiratory causes was insensitive to the 

7 use of a random-effects survival model allowing for spatial clustering within the 

8 metropolitan area and state of residence, and adjustment for several ecologic variables 

9 considered individually. Subgroup analyses showed that temperature and region of 

10 country, but not sex, age at enrollment, body-mass index, education, or PM2.5

11 concentration, modified the effects of O3 on the risk of death from respiratory causes 

12 (i.e., risks were higher at higher temperature, and in the Southeast, Southwest, and Upper 

13 Midwest). Ozone threshold analyses indicated that the threshold model was not a better 

14 fit to the data (p > 0.05) than a linear representation of the overall O3-mortality 

15 association. Overall, this new analysis indicates that long-term exposure to PM2.5

16 increases risk of cardiac death, while long-term exposure to O3 is specifically associated 

17 with an increased risk of respiratory death, and suggests that combining cardiovascular 

18 and respiratory causes of mortality into one category for analysis may obscure any effect 

19 that O3 may have on respiratory-related causes of mortality.

Table 7-12 Relative risk (and 95% CI) of death attributable to a 10-ppb change 

in the ambient O3 concentration* 

Cause of Death O3 (96 MSAs) O3 (86 MSAs) O3 +PM2.5 (86 MSAs)

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)

* 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 per cubic meter were recorded for members of the cohort in 1999 and 2000.

Source: Reprinted with permission of Massachusetts Medical Society (Jerrett et al., 2009)

20 In a similar analysis, Smith et al. (2009a) used data from 66 MSAs in the ACS cohort to 

21 examine the association of O3 concentrations during the warm season and all-cause and 

22 cardiopulmonary mortality. Mortality effects were estimated in single pollutant and 

23 copollutant models, adjusting for two PM2.5 constituents, sulfate and EC. When all-cause 

24 mortality was investigated, there was a 0.8% (95% CI: -0.31, 1.9) increase associated 

25 with a 10 ppb increase in O3 concentration. This association was diminished when sulfate 

26 or EC were included in the model. There was a 2.48% (95% CI: 0.74, 4.3) increase in 

27 cardiopulmonary mortality associated with a 10 ppb increase in O3 concentration. The 
Draft – Do Not Cite or Quote 7-85 September 2011

1 cardiopulmonary association was robust to adjustment for sulfate, and diminished, though 

2 still positive, after adjustment for EC (1.63% increase; 95% CI: -0.41, 3.7). Smith et al. 

3 (2009a) did not specifically separate out cardiovascular and respiratory causes of death 

4 from the cardiopulmonary category, as was done by Jerrett et al. (2009).

7.7.1 Summary and Causal Determination

5 The 2006 O3 AQCD concluded that an insufficient amount of evidence existed “to 

6 suggest a causal relationship between chronic O3 exposure and increased risk for 

7 mortality in humans” (U.S. EPA, 2006b). Several additional studies have been conducted 

8 since the last review, including an ecologic study that finds no association between 

9 mortality and O3 (Wang et al., 2009c), a study of four cohorts of Medicare enrollees with 

10 potentially predisposing conditions that observes associations between O3 and mortality 

11 among each of the cohorts (Zanobetti and Schwartz, In Press), and reanalyses of the ACS 

12 cohort that provide weak evidence for an association with cardiopulmonary mortality 

13 (Smith et al., 2009a) and specifically point to a relationship between long-term O3

14 exposure and an increased risk of respiratory mortality (Jerrett et al., 2009). The findings 

15 from the Jerrett et al. (2009) study are consistent and coherent with the evidence from 

16 epidemiologic, controlled human exposure, and animal toxicological studies for the 

17 effects of short- and long-term exposure to O3 on respiratory effects. Additionally, the 

18 evidence for short- and long-term respiratory morbidity provides biological plausibility 

19 for mortality due to respiratory disease. Collectively, the evidence is suggestive of a 

20 causal relationship between long-term O3 exposures and mortality.

7.8 Overall Summary

21 The evidence reviewed in this chapter describes the recent findings regarding the health 

22 effects of long-term exposure to ambient O3 concentrations. Table 7-13 provides an 

23 overview of the causal determinations for each of the health categories evaluated.
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Table 7-13 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

Mortality Suggestive of a causal relationship

7.9 References

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inhalable particles and other air pollutants related to mortality in nonsmokers. Am J Respir Crit Care 

Med 159: 373-382. 

Akinbami, LJ; Lynch, CD; Parker, JD; Woodruff, TJ. (2010). The association between childhood asthma 

prevalence and monitored air pollutants in metropolitan areas, United States, 2001-2004. Environ Res 

110: 294-301. http://dx.doi.org/10.1016/j.envres.2010.01.001.

Angoa-Pérez, M; Jiang, H; Rodríguez, AI; Lemini, C; Levine, RA; Rivas-Arancibia, S. (2006). Estrogen 

counteracts ozone-induced oxidative stress and nigral neuronal death. Neuroreport 17: 629-633. 

Arito, H; Uchiyama, I; Yokoyama, E. (1992). Acute effects of ozone on EEG activity, sleep-wakefulness and 

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