Chapter 7. Integrated Health Effects of

Long-Term Ozone Exposure
 

7.1. In tro d u c tio n 

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

2 effects and long-term exposures to O3. Both epidemiologic and toxicological studies provide a basis 

3 for examining long-term O3 exposure health effects for respiratory effects, cardiovascular effects,
 

4 reproductive and developmental effects, central nervous system effects, cancer outcomes, and 

5 mortality.
 

6 Conclusions from the 2006 O3 AQCD are summarized briefly at the beginning of each section,
 

7 and the evaluation of evidence from recent studies builds upon what was available during the 

8 previous review. For each health outcome (e.g., respiratory disease, lung function), results are 

9 summarized for studies from the specific scientific discipline, i.e., epidemiologic and toxicological 

10 studies. The major sections (i.e. respiratory, cardiovascular, mortality, reproductive/developmental, 

11 cancer) conclude with summaries of the evidence for the various health outcomes within that 

12 category and integration of the findings that lead to conclusions regarding causality based upon the 

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

14 category, such as respiratory effects, with coherence and plausibility being based on evidence from 

15 across disciplines and also across the suite of related health outcomes, including cause-specific 

16 mortality. 

7.2. Res piratory Effects 

17 Studies reviewed in the 2006 O3 AQCD (U.S. EPA, 2006, 088089) examined evidence for 

18 relationships between long-term O3 exposure and effects on respiratory health outcomes including 

19 seasonal declines in lung function, increases in inflammation, and development of asthma in children 

20 and adults. The term seasonal was used in these studies as a measure of a long-term exposure of 

21 several months. Animal toxicology data provided a clearer picture indicating that long-term O3 

22 exposure may have lasting effects. Chronic exposure studies in animals have reported biochemical 

23 and morphological changes suggestive of irreversible long-term O3 impacts on the lung. In contrast 

24 to supportive evidence from chronic animal studies, the epidemiologic studies on longer-term lung 

25 function declines, inflammation, and new asthma development remained inconclusive. Several 

26 studies (e.g., Frischer et al., 1999, 001037; Horak et al., 2002, 034792) collectively indicated that 

Note: Hyperlinks to the reference citations throughout this document will take you to the NCEA HERO database (Health and 

Environmental Research Online) at http://epa.gov/hero. HERO is a database of scientific literature used by U.S. EPA in the process of 

developing science assessments such as the Integrated Science Assessments (ISAs) and the Integrated Risk Information System (IRIS). 

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1 seasonal O3 exposure was associated with smaller increases in lung function growth in children. For 

2 longer time periods, the definitive analysis in the Child Health Study (CHS) reported by Gauderman 

3 et al. (2004, 056569) provided little evidence that long-term exposure to ambient O3 at current levels 

4 was associated with significant deficits in the growth rate of lung function in children in contrast to 

the effects observed with other pollutants such as acid vapor, NO2, and PM2.5. Asthmatic children 

6 with GSTM1 null genotype were found to be more susceptible to the impact of O3 exposure on small 

7 airways function in Mexico (Romieu et al., 2004, 056796). Limited epidemiologic research 

8 examined the relationship between long-term O3 exposures and inflammation. Inflammatory effects 

9 consistent with known effects of O3 such as increased eosinophil levels were observed in an Austrian 

study (Frischer et al., 2001, 019683). The cross-sectional surveys available for the 2006 O3 AQCD 

11 detected no associations between long-term O3 exposures and asthma prevalence, asthma-related 

12 symptoms or allergy to common aeroallergens in children after controlling for covariates. 

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

14 and new-onset asthma related to genotype in U.S. cohorts in multi-community studies. Related 

studies report coherent relationships between respiratory symptoms among asthmatics and long-term 

16 O3 exposure. A new line of evidence reports a positive exposure response relationship between first 

17 asthma hospitalization and long-term O3 exposure. Results from recent studies examining pulmonary 

18 function, inflammation, and allergic responses are also presented. 

7.2.1. Ne w Ons e t As th ma 

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

environmental exposure (Gilliland et al., 1999, 155792). Complex chronic diseases, such as asthma, 

21 are partially the result of a sequence of biochemical reactions involving exposures to various 

22 environmental agents metabolized by a number of different genes (Conti et al., 2003, 626696). 

23 Understanding the relation between genetic polymorphisms and environmental exposure can help 

24 identify high-risk subgroups in the population and provide better insight into pathway mechanisms 

for these complex diseases. Oxidative stress likely underlies these mechanistic hypotheses (Gilliland 

26 et al., 1999, 155792). Susceptibility genes act through modification of disease risk associated with 

27 environmental factors. Epidemiologic investigation of hypotheses of possible mechanisms involving 

28 the gene-environmental (GxE) interaction involves statistical analysis of these interactions for the 

29 risk of new-onset asthma in children being influenced by exposure to air pollution (Gauderman, 

2001, 625862; Gauderman, 2002, 626945; Gilliland et al., 1999, 155792). 

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

32 onset asthma are provided by several recent studies (Ercan et al., 2006, 595172; Gilliland et al., 

33 2002, 090970; Hanene et al., 2007, 595428; Himes et al., 2009, 480112; Islam et al., 2008, 097348; 

34 Li et al., 2006, 596447; Li et al., 2008, 596449; Tamer et al., 2004, 199914). Evidence for a genepollution interaction in the pathogenesis of asthma are supported by recent study findings (Gilliland 

36 et al., 2002, 090970; Islam et al., 2008, 097348; Islam et al., 2009, 196715; Lee et al., 2004, 090971; 

37 Oryszczyn et al., 2007, 596460). 

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Evidence for associations between long-term exposureto O3 and new-onset asthma is provided 

by new studies from the CHS. Initiated in the early 1990's, the CHS was originally designed to 

examine whether ambient pollutants were related to chronic respiratory outcomes in children (Peters 

et al., 1999, 087243; Peters et al., 1999, 087237). About 10 years later, the CHS inaugurated a series 

of genetic studies (Gilliland et al., 1999, 155792) nested within the CHS cohort by obtaining 

biological samples from the study subjects (buccal cells). These new studies examined the 

relationship between health outcomes, genetic susceptibility, behavioral factors and environmental 

exposure. 

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

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

evaluated, and then whether the effects of these variants varied by exposure to O3 (Islam et al., 2008, 

097348). HMOX1 [heme oxygenase (decycling) 1] is a human gene that encodes for the enzyme 

heme oxygenase. Heme oxygenase 1 (HO-1) is an enzyme that catalyzes the metabolism of heme. 

The heme iron serves as a source or sink of electrons during electron transfer or redox chemistry, so 

the presence of the HMOX1 gene, and therefore the generation of heme oxygenase, protects against 

oxidative stress in the body. The authors observed that functional promoter variants in CAT and 

HMOX-1 showed ethnicity-specific associations with new-onset asthma and that oxidant gene 

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

The subjects were obtained from the CHS from 12 communities in southern California. 

Children with a history of asthma or wheeze were excluded from this analysis. Analyses were 

restricted to children of Hispanic (n = 576) or non-Hispanic white ethnicity (n = 1,125). New-onset 

asthma was classified as such for children with no prior history of asthma at study entry who 

subsequently reported physician-diagnosed asthma at annual follow-up with the date of onset 

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

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

restricted to those new-onset asthma cases who also used an inhaler (n = 121). Long-term pollutant 

levels were calculated from 1994 through 2003. The effect of ambient air pollution on the 

relationship between genetic polymorphism and new-onset asthma was assessed using models where 

the community specific average air pollution levels were fitted as a continuous variable together with 

the appropriate interaction terms for genes and air pollutants (Berhane et al., 2004, 626732). Cox 

proportional hazard regression models were fitted to the data. A stratified analysis for the two 

independent fourth-grade cohorts of the study population recruited in 1993 and 1996 were conducted 

to assess whether the results could be replicated in independent groups of children. 

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

097348). The evidence indicated that the effect of variation in the HMOX-1 gene on risk of newonset asthma differed by ambient O3 level. An interaction P value was reported of 0.003 from the 

hierarchical two stage Cox proportional hazard model fitting the community-specific O3 and PM10 

levels (continuous) and controlling for random effect of the communities. The annual O3 levels 

(10:00 a.m. – 6:00 p.m.) ranged from 46.5 to 64.9 ppb in the six higher O3 communities (mean = 

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55.2 ppb) and 28.6 to 45.5 ppb in the six lower O3 communities (mean = 38.4 ppb). Average O3 

levels showed low correlation with the other monitored pollutants. The interaction indicated a greater 

effect (association) of community O3 level among children with the gene than with children without 

the gene. Alleles with 23 or fewer (GT)n repeats are categorized as short (S). The S-allele variant of 

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

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

were S-allele carriers and resided in low-O3 communities with Hazard Ratio (HR) of 0.44 (95% CI: 

0.23, 0.83). The ratio of HR of S-allele carriers who resided in high O3 communities (HR 0.88; [95% 

CI: 0.33, 2.34]) was twofold greater than in those who resided in the low-O3 communities (HR 0.44). 

The non-parallelism of the two lines in Figure 7-1 illustrates the interaction: Children with the Sallele have protection against the onset of asthma; however, in high- O3 communities, this protection 

is attenuated. The results from sensitivity analyses on the two fourth-grade cohorts, and the inhaler 

definition for asthma were both consistent with the main results. An analysis related to children’s 

participation in sports or time spent outdoors produced the same outcome. No significant interactions 

were observed between PM10 or other pollutants and the HMOX -1 gene. A potential concern for not 

adjusting for multiple testing was considered by the authors as not a factor in this analysis because 

the selection of the genes was based on a priori hypotheses defined by a well-studied biological 

pathway. Thus in this cohort in southern California, Islam et al. (2008, 097348) related new-onset 

asthma to O3 exposure in genetically susceptible children. 

Related to the findings in Islam et al. (2008, 097348) discussed above, Islam et al. (2009, 

196715) examined whether GSTP1, GSTM1, exercise and O3 exposure have interrelated effects on 

the pathogenesis of asthma. A modifying role of air pollution on the association between Ile105Val 

and asthma in a cohort of children had been observed (Lee et al., 2004, 090971), but the study did 

not examine O3 specifically or consider exercise. A primary conclusion that the authors (Islam et al., 

2009, 196715) reported was that the common functional variants of GSTP1 and GSTM1 null 

genotypes modulate the risk of new onset asthma during adolescence. Children who had the GSTM1 

null genotype were at 1.6-fold (95% CI: 1.2, 2.2) increased risk of developing new onset asthma 

compared with those without the null genotype. Further, the CHS investigators examined the 

complex interrelationship of antioxidant defenses with asthma risk with increasing doses of O3, 

resulting from increasing ventilation associated with vigorous exercise characterized by the number 

of team sports played. In an earlier analysis, McConnell et al. (2002, 023150) had reported that the 

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

but did not consider genetic variants. The plausibility of a causal association is strengthen by the 

observation by Islam et al. (2009, 196715) that the risk of participation in team sports was related to 

increased genetic susceptibility to oxidative stress. The sixfold increased risk of asthma (HR 6.15, 

[95% CI: 2.2, 7.4]) for children who were homozygous for Ile105, participated in three or more team 

sports and lived in high-O3 communities demonstrates the potential importance of a combination of 

genetic variability, O3 exposure and behavior on asthma risk. 

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Interaction of Gene presence and Ozone Level on the 

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

0.94 0.88 

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0.44 

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0.5 

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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) 

Source:Used with permission from American Thoracic Society, Islam et al. (2008, 097348). 

Figure 7-1. Interaction of gene presence and ozone level on the Hazard Ratio (HR) of new-onset

asthma in the 12 Children's Health Study communities. ﷒An interaction P-value of 

0.003 was obtained from the hierarchical two stage Cox proportional hazard model

fitting the community specific ozone and controlling for random effect of the

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

community ozone 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. 

1 Epidemiologic evidence of associations of arginase variants with asthma are limited (Li et al., 

2 2006, 596447). Asthmatic subjects have higher arginase activity than nonasthmatic subjects (Morris 

3 et al., 2004, 674145). NO is a mediator of nitrosative stress synthesized from L-arginine by nitric 

4 oxide synthases. In the CHS, Salam et al. (2009, 596644) examined whether arginase variants 

5 (ARG1 and ARG2 genes) were associated with asthma and whether atopy and exposures to smoking 

6 and air pollution influence the associations. The modifying effect of O3 and atopy on the association 

7 between haplotypes and asthma were evaluated using likelihood ratio tests with appropriate 

8 interaction terms. They found that both ARG1 and ARG2 genetic loci were associated with 

9 childhood-onset asthma. The effect of the ARG1 haplotype varied by the child’s history of atopy and 

10 ambient O3. Among atopic children living in high O3 communities, those carrying the ARG1 

11 halotype had reduced asthma risk (OR per ARG1h4 halotype copy: 0.12; [95% CI: 0.04, 0.43]; 

12 P heterogeneity across atopy/O3 categories = 0.008). 

13 Further, the CHS presents results examining the relationship of new onset asthma with traffic14 related pollution near homes and schools (McConnell et al., 2010, 625501). Asthma risk increased 

15 with modeled traffic-related pollution exposure from roadways near homes and near schools. The 

16 HR was 0.76 (95% CI: 0.38, 1.54) across the range of ambient O3 exposure in the communities. With 

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adjustment for school and residential non-freeway traffic-related exposure, the estimated HR for O3 

was 1.01 (95% CI: 0.49, 2.11). Gene variants were not evaluated in this study. 

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

between chronic exposure to O3 and prevalence of asthma and asthmatic symptoms in school 

children (Ramadour et al., 2000, 013259; Wang et al., 1999, 008105) while others (Charpin et al., 

1999, 015152; Kuo et al., 2002, 036310) did not. Recent studies provide additional evidence. 

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

(2005, 089454) assessed the effects of air pollutants on the risk of asthma. The study population was 

recruited from elementary and middle schools within 1 km of air monitoring stations. The risk of 

asthma was related to O3 in the one-pollutant model. The addition of other pollutants, in twopollutant and three-pollutant models, increased the O3 risk estimates. The prevalence of childhood 

asthma was assessed in Portugal by contrasting the risk of asthma between a high O3 rural area and 

an area with low O3 levels (Sousa et al., 2008, 619959; Sousa et al., 2009, 619956; Sousa et al., 

2011, 676712). The locations were selected to provide a difference in O3 levels without the 

confounding effects of other pollutants. Both evaluation for asthma symptoms and FEV1 suggested 

that O3 increased asthma prevalence. Clark et al. (2010, 594440) investigated the effect of exposure 

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

asthma diagnosis up to 3-4 years of age in a population-based nested case-control study for all 

children born in southwestern British Columbia in 1999 and 2000 (n=37,401; including 3,482 [9.3%] 

with asthma). Air pollution exposure for each subject was estimated based on their residential 

address history using regulatory monitoring data, land use regression modeling, and proximity to 

stationary pollutant sources. Daily values from the three closest monitors within 50 km were used to 

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

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

asthma diagnosis in infants makes this study difficult to interpret (Martinez et al., 1995, 046150). In 

a cross-sectional analysis, Akinbamia et al. (2010, 378580) examined the association between 

chronic exposure to outdoor pollutants (12-month average levels by county) and asthma outcomes in 

a national sample of children ages 3-17 years living in U.S. metropolitan areas (National Health 

Interview Survey, N = 34,073). A 5-ppb increase in estimated 8-h max O3 concentration (annual 

average) yielded a positive association for both currently having asthma and for having at least 1 

asthma attack in the previous year; while the adjusted odds ratios for other pollutants were not 

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

comparable results. Multi-pollutant models produced similar results. The median value for 12-month 

average O3 levels was 39.5 ppb and the IQR was 35.9-43.7 ppb. The adjusted odds for current 

asthma for the highest quartile (49.9-59.5 ppb) of estimated O3 exposure was 1.56 (95% CI: 1.15, 

2.10) with a positive dose-response relationship apparent from the lowest quartile to the highest. 

Thus, this cross-sectional analysis and Hwang et al. (2005, 089454) provides further evidence 

relating O3 exposure and the risk of asthma. 

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

2 CHS examining the role of gene-environment interactions and long-term O3 exposure. Lee et al. 

3 (2009, 199915) studied associations of TNF-308 genotype with bronchitis symptoms among 

4 asthmatic children and investigated whether associations vary with ambient O3 exposure since 

5 increased airway TNF may be related to inflammation. Asthmatic children with the GG genotype 

6 had a lower prevalence of bronchitic symptoms compared with children carrying at least one A-allele 

7 (e.g., GA or AA). Low-versus high-O3 strata were defined as less than or greater than 50- ppb O3 

8 average. Asthmatic children with TNF-308 GG genotype had a significantly reduced risk of 

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

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

11 in genotypic effects between low- and high-O3 environments was statistically significant among 

12 asthmatics (P for interaction = 0.01), but insignificant among non-asthmatic children. By using 

13 dummy variables in each community, Lee et al. (2009, 199915) calculated the effect of TNF-308 GG 

14 genotype on the occurrence of bronchitic symptoms among children with asthma. Figure 7-2 

15 presents adjusted O3 community-specific beta-coefficients plotted against ambient O3 concentration, 

16 using weights proportional to the inverse variance. They further report that they found no substantial 

17 differences in the effect of the GG genotype in asthmatic children in relation to exposure to PM10, 

18 PM2.5, NO2, acid vapor or second-hand smoke exposure. These results suggest a role of gene19 environment interactions such as long-term O3 exposure on the occurrence of bronchitic symptoms 

20 among children with asthma. 

Source: Used with permission from John Wiley & Sons A/S, Lee et al. (2009, 199915). 

Figure 7-2. Ozone modifies the effect of TNF G-308A genotype on bronchitic symptoms among 

children with asthma in the CHS. ﷒Using dummy variables in each of the 12 

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

bronchitic symptoms among children with asthma. 

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

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

French cities (Paris, Lyon, Marseille, Montpellier, and Grenoble) (Rage et al., 2009, 196720). In this 

cross-sectional study, asthma severity over the past 12 months was assessed among 328 adult 

asthmatics using two methods: (1) a four-class severity score that integrated clinical events and type 

of treatment; and (2) a five-level asthma score based only on symptoms. Two measures of exposure 

were also assessed: (1 [first method]) closest monitor data from 1991 to 1995 where a total of 93% 

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

interpolations and then assigned air pollutants to the geocoded residential addresses of all 

participants and individually assigned exposure to ambient air pollution estimates. Higher asthma 

severity scores were significantly related to both the 8-h avg O3 during April-September and the 

number of days with 8-h O3 averages above 55 ppb. Both exposure assessment methods and severity 

score methods resulted in very similar findings. Effect estimates of O3 were similar in three-pollutant 

models. No PM data were available. Since these estimates were not sensitive to the inclusion of 

ambient NO2 in the three-pollutant models, the authors viewed the findings not to be explained by 

particles which usually have substantial correlations between PM and NO2. Ozone concentrations by 

the first method for annual levels, nearest monitor were 8-hours (n = 210); mean (+SD) 30.25 (+9.7); 

IQR: 21-36.5 (+15.5) ppb. The second exposure approach yielded summer (n = 308); mean (+SD) 

levels of 31.5 (+5.2); IQR of 28.5-33.9 (+5.5) ppb. Effect estimates for O3 in three-pollutant models 

including O3, SO2, and NO2 yielded OR for O3-days of 2.74 (95% CI: 1.68, 4.48) per IQR days of 

10-28 (+18) ppb. The effect estimates for SO2 and NO2 in the three-pollutant model were 1.33 (95% 

CI: 0.85, 2.11) and 0.94 (95% CI: 0.68, 1.29) respectfully. Taking into account duration of residence 

did not change the result. This study suggests that a higher asthma severity score is related to longterm O3 exposure. 

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

ambient pollutants, and acute respiratory illness were investigated in a national U.S. cohort (Wenten 

et al., 2009, 597084). Health information, air pollution, and incident respiratory-related school 

absences were ascertained in January-June 1996 for 1,136 Hispanic and non-Hispanic white U.S. 

elementary schoolchildren as part of the prospective Air Pollution and Absence Study, a population 

based cohort study conducted as part of the CHS. A related earlier study (Gilliland et al., 2001, 

013232), which was discussed in the 2006 O3 AQCD, examined the effects of ambient air pollution 

on school absenteeism due to respiratory illness without a genetic aspect to the study. In a new study 

Wenten et al. (2009, 597084) hypothesized that variation in the level or function of these enzymes 

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

effect of these two genes on respiratory illness was examined. Risk of respiratory-related school 

absences was elevated for children with the CAT (G/G) and MPO (G/A or A/A) genes (relative risk = 

1.35, [95% CI: 1.03, 1.77]; P-interaction = 0.005). To assess effects of long-term average levels of 

O3 on acute effects, communities were divided into high and low exposure groups by median levels 

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1 (46.9 ppb O3). The epistatic effect of CAT and MPO variants was evident in communities exhibiting 

2 high ambient O3 levels (P-interaction = 0.03). The association of respiratory-illness absences with 

3 functional variants in CAT and MPO that differ by air pollution levels illustrates the need to consider 

4 genetic epistasis in assessing gene-environment interactions. In high O3 communities, CAT/MPO 

genotypes that resulted in decreased oxidative stress were associated with a decreased risk of 

6 respiratory related school absences compared with the CAT/MPO wild-type genotype (RR = 0.42, 

7 [95% CI: 0.20, 0.89]). 

7.2.2. As thma Hos pital Admis s ions and ED Vis its 

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

9 asthma and respiratory disease that were available in the 2006 O3 AQCD mainly looked at the daily 

time metric. New studies evaluated long-term O3 exposure metrics providing a new line of evidence 

11 that suggests a positive exposure-response relationship between first asthma hospital admission and 

12 long-term O3 exposure. 

13 An ecologic study (Moore et al., 2008, 196685) evaluated time trends in associations between 

14 declining warm-season O3 concentrations and hospitalization for asthma in children in California’s 

South Coast Air Basin who ranged in age from birth to 19 years. Quarterly average concentrations 

16 from 195 spatial grids, 10×10 km, were used. Ozone was the only pollutant associated with 

17 increased hospital admissions over the study period. A linear relation was observed for asthma 

18 hospital discharges (Moore et al., 2008, 196685). A matched case-control study (Karr et al., 2007, 

19 090719) was conducted of infant bronchiolitis (ICD 9, code 466.1) hospitalization and two measures 

of long-term pollutant exposure (the month prior to hospitalization and the lifetime average) for O3 

21 in the South Coast Air Basin of southern California among 18,595 infants born between 1995 and 

22 2000. Ozone was associated with reduced risk in the single-pollutant model, but this relation did not 

23 persist in multi-pollutant models. 

24 In a cross-sectional study, Meng et al. (2010, 594252) examined associations between air 

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

26 Health Interview Survey data from subjects ages 1 to 65+ who reported physician-diagnosed asthma 

27 (n = 1502). Subjects were assigned annual average concentrations for O3 based on residential ZIP 

28 code and the closet air monitoring station within 8 km but did not have data on duration of residence. 

29 Multi-pollutant models for O3 and PM did not differ substantially from single-pollutant estimates, 

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

31 increased asthma-related ED visits or hospitalizations for O3 (OR 1.49; [95% CI: 1.05, 2.11] per 

32 10 ppb) for all ages. Positive associations were obtained for symptoms but 95% CIs included null 

33 values. Associations for symptoms for adults (ages 18 +) were observed (OR 1.40; [95% CI: 1.02, 

34 1.91} per 10 ppb). 

Associations between air pollution and poorly controlled asthma among adults in Los Angeles 

36 and San Diego Counties, were investigated using the California Health Interview Survey data 

37 collected between November 2000 and September 2001 (Meng et al., 2007, 093275). Each 

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respondent was assigned an annual average concentration measured at the nearest station within 

5 miles of the residential cross-street intersection. Poorly controlled asthma was defined as having 

daily or weekly asthma symptoms or at least one ED visit or hospitalization because of asthma 

during the past 12 months. This cross-sectional study reports an OR of 3.34 (95% CI: 1.01, 11.09) 

for poorly controlled asthma when comparing those 65 years of age and older above the 90th 

percentile (28.7 ppb) level to those below that level. Multi-pollutant analysis produced similar 

results. 

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

concentration-response relationship is provided in a retrospective cohort study (Lin et al., 2008, 

196680). This study investigated the association between chronic exposure to O3 and childhood 

asthma admissions (defined as a principal diagnosis of ICD9, code 493) by following a birth cohort 

of 1,204,396 eligible births born in New Yor k State during 1995-1999 to first asthma admission or 

until 31 December 2000. There were 10,429 (0.87%) children admitted to the hospital for asthma 

between 1 and 6 years of age. The asthma hospitalization rate in New Yor k State in 1993 was 2.87 

per 1,000 (Lin et al., 1999, 377917). Three indicators (all 8-h max from 10:00 a.m. to 6:00 p.m.) 

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

(41.06 ppb); (2) mean concentration during the O3 season (50.62 ppb); and (3) proportion of followup days with O3 levels >70 ppb. In this study the authors aimed to predict the risk of having asthma 

admissions in a birth cohort, but the time to the first admission in children that is usually analyzed in 

survival models was not their primary interest. The effects of co-pollutants were assessed and 

controlled for using the Air Quality Index (AQI). Interaction terms were used to assess potential 

effect modifications. A positive association between chronic exposure to O3 and childhood asthma 

hospital admissions was observed indicating that children exposed to high O3 levels over time are 

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

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

Medicaid/self-paid births, geographic region and others. As shown in Figure 7-3, positive 

concentration-response relationships were observed. Asthma admissions were significantly 

associated with increased O3 levels for all chronic exposure indicators (ORs, 1.16-1.68). When 

estimating the O3 effect using the exceedance proportion, an increase was observed (OR 1.68; [95% 

CI: 1.64, 1.73]) in hospital admissions with an IQR (2.51%) increase in O3. A proportional hazards 

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

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

CI: 1.124, 1.155] is similar to OR from the logistic model: 1.16 (95% CI: 1.15, 1.17) (Lin, personal 

communication, 2010, 676733). Thus, this study provides evidence associating long-term O3 

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

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Source: Lin et al. (2008, 196680);(Lin, personal communication, 2010, 676733). 

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

during the entire follow-up period 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. 

OR’s 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. 

7.2.3. Pulmonary Structure and Function 

1 The definitive 8-year follow-up analysis of the first cohort of the CHS (Gauderman et al., 

2 2004, 056569) provided little evidence that long-term exposure to ambient O3 at current levels was 

3 associated with significant deficits in the growth rate of lung function in children. A later CHS study 

4 (Islam et al., 2007, 090697) examined relationships between air pollution, lung function, and new 

5 onset asthma and reported no substantial differences in the effect of lung function between “high-” 

6 and “low-” O3 communities. Ozone concentrations from the least to most polluted communities 

7 (mean annual average of 8-h avg O3) ranged from 30 to 65 ppb, whereas the ranges observed for the 

8 other pollutants had four- to eightfold differences in concentrations. In a more recent CHS study, 

9 Breton et al. (2011, 687660) hypothesized that genetic variation in genes on the glutathione 

10 metabolic pathway may influence the association between ambient air pollutant exposures and lung 

11 function growth in children. They investigated whether genetic variation in glutathione genes GSS, 

12 GSR, GCLC, and GCLM was associated with lung function growth in healthy children using data 

13 collected on 2,106 children over an 8-year time-period as part of the Children’s Health Study. Breton 

14 et al. (2011, 687660) found that variation in the GSS locus was associated with differences in 

15 susceptibility of children for lung function growth deficits associated with NO2, PM10, PM2.5, 

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elemental carbon, organic carbon, and O3. The negative effects of air pollutants were largely 

observed within participants who had a particular GSS haplotype. The effects ranged from -124.2 to 

-149.1 mL for FEV1, -92.9 to -126.7 mL for FVC and -193.9 to -277.9 mL/s for MMEF for all 

pollutants except O3, for which some positive associations were reported: 25.9 mL for FEV1; 0.1 mL 

for FVC, and 166.5 mL/s for MMEF. Ozone did show larger decreases in lung function in children 

without this haplotype, when compared to the other pollutants with values of -76.6 mL for FEV1, 

-17.2 mL for FVC, and -200.3 mL/s for MMEF, but only MMEF was statistically significant. 

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

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

small airways (<2 mm) function (FEF75 and FEF25-75) (Tager et al., 2005, 087538). There was an 

interaction with the FEF25-75/FVC ratio, a measure of intrinsic airway size. Subjects with a large ratio 

were less likely to have decreases in FEF75 and FEF25-75 for a given estimated lifetime exposure to 

O3. Kinney and Lippmann (2000, 011913) examined 72 nonsmoking adults (mean age 20 years) 

from the second-year class of students at the U.S. Military Academy in West Point, NY, and reported 

results that appear to be consistent with a seasonal decline in lung function that may in part be due to 

O3 exposures. Ilhorst et al. (2004, 055608) examined 2,153 children with a median age of 7.6 years 

and reported summer pulmonary function results which indicated that significantly lower FVC and 

FEV1 increases were associated with higher O3 exposures in the summer, but not in the winter. Semiannual mean O3 concentrations ranged from 22 to 54 ppb during the summer and 4 to 36 ppb during 

the winter. However, over a 3.5-year period Ilhorst et al. (2004, 055608) found no associations 

between increases in lung function and mean summer O3 levels for FVC and FEV1, in contrast to the 

significant seasonal effects. Frischer et al. (1999, 001037) showed results similar to the Ilhorst et al. 

(2004, 055608) study. 

Mortimer et al. (2008, 122163; 2008, 187280) examined the association of prenatal and 

lifetime exposures to air pollutants with pulmonary function and allergen sensitization in a subset of 

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

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

separately across several important developmental time-periods, including: the entire pregnancy, 

each trimester, the first 3 years of life, the first 6 years of life, and the entire lifetime. In the first 

analysis (Mortimer et al., 2008, 122163), negative effects on pulmonary function were found for 

exposure to PM10, NO2, and CO during key neonatal and early life developmental periods. The 

authors did not find a negative effect of exposure to O3 within this cohort. In the second analysis 

(Mortimer et al., 2008, 187280), sensitization to at least one allergen was associated, in general, with 

higher levels of CO and PM10 during the entire pregnancy and second trimester, and higher PM10 

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

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

time periods were correlated, the strongest associations with the outcomes were observed for 

prenatal exposures. Though it may be difficult to disentangle the effect of prenatal and postnatal 

exposures, the models from this group of studies suggest that each time period of exposure may 

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contribute independently to different dimensions of school-aged children’s pulmonary function. For 

4 of the 8 pulmonary-function measures (FVC, FEV1, PEF, FEF25-75), prenatal exposures were more 

influential on pulmonary function than early-lifetime metrics, while, in contrast, the ratio of 

measures (FEV1/FVC and FEF25-75/FVC) were most influenced by postnatal exposures. When 

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

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

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

for lifetime or current exposures. 

Latzin et al. (2009, 195721) examined whether prenatal exposure to air pollution was 

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

inhomogeneity and eNO were measured in 241 unsedated, sleeping neonates (age = 5 weeks). 

Consistent with the previous studies, no association was found for prenatal exposure to O3 and lung 

function. 

In a cross-sectional study of adults, Qian et al. (2005, 093283) examined the association of 

long-term exposure to O3 and PM10 with pulmonary function from data of 10,240 middle-aged 

subjects who participated in the Atherosclerosis Risk in Communities (ARIC) study in four U.S. 

communities. A surrogate for long-term O3 exposure from daily data was determined at the 

individual level. Ozone was significantly and negatively associated with measures of pulmonary 

function. 

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

decline in lung function, Forbes et al. (2009, 595425) studied the association between chronic 

exposure to outdoor air pollution and lung function in approximately 42,000 adults aged 16 and 

older who were representatively sampled cross-sectionally from participants in the Health Survey for 

England (1995, 1996, 1997, and 2001). FEV1 was not associated with O3 concentrations. In contrast 

to the results for PM10, NO2, and SO2; combining the results of all the survey years showed that a 

5-ppb difference in O3 was counter-intuitively associated with a higher FEV1 by 22 mL. 

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

(n = 3,170) who were 8 years of age at the beginning of the study, Rojas-Martinez et al. (2007, 

091064) evaluated the association between long-term exposure to O3, PM10 and NO2 and lung 

function growth every 6 months from April 1996 through May 1999. Exposure data were provided 

by 10 air quality monitor stations located within 2 km of each child’s school. Over the study period, 

8-h O3 concentrations ranged from 60 ppb (SD, ±25) in the northeast area of Mexico City to 90 ppb 

(SD, ±34) in the southwest, with an overall mean of 69.8 ppb. In multi-pollutant models, an IQR 

increase in mean O3 concentration of 11.3 ppb was associated with an annual deficit in FEV1 of 

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

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

co-pollutant models were not substantially different than single pollutant models, independent effects 

for pollutants could not be estimated accurately because the traffic-related pollutants were correlated. 

To reduce exposure misclassification, microenvironmental and personal exposure assessments were 

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1 conducted in a randomly selected subsample of 60 children using passive O3 samplers. Ozone 

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

3 monitoring stations. 

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

pulmonary function. The strongest evidence was for seasonal effects of extended O3 exposures on 

6 lung function in children, i.e., reduced lung function growth being associated with higher ambient O3 

7 levels. Longer-term studies, investigating the association of chronic O3 exposure on yearly lung 

8 function such as the CHS, were inconclusive. Thus for new studies for pulmonary function, in one 

9 study where O3 and other pollutant levels were higher (90 ppb at high end of the range) than those in 

the CHS, a relationship between O3 concentration and pulmonary function declines was observed in 

11 school-aged children. Two studies of adult cohorts provide mixed results where long-term exposures 

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

13 Thus there is little new evidence to build upon the very limited studies from the 2006 O3 AQCD. 

7.2.3.1. Evidence from Toxicological Studies 

14	 As reviewed in the 1996 and 2006 O3 AQCDs (U.S. EPA, (1996, 017831), (2006, 088089)), 

considerable controversy surrounds the extrapolation of data generated by rodent toxicology studies 

16 to the understanding of adverse health effects observed in humans, as documented by epidemiology 

17 and controlled exposure studies. Chief among these data extrapolation issues are the differences 

18 between rodent and human respiratory physiology, cellular makeup, dosimetry, and morphometry. 

19 Unique among the six NAAQS criteria pollutants, however, O3-inhalation studies have been 

performed in non-human primates whose respiratory system most closely resembles that of the 

21 human. A long series of studies have used non-human primates to examine the effect of O3 alone or 

22 in combination with an inhaled allergen, house dust mite antigen, on morphology and lung function. 

23 These studies, by Plopper and colleagues, have demonstrated changes in pulmonary function and 

24 airway morphology in adult and infant non-human primates repeatedly exposed to environmentally 

relevant concentrations of O3. (Carey et al., 2007, 195752; Chang MM-J; Wu et al., 1998, 011983; 

26 Chen et al., 2003, 035576; Duan et al., 1993, 086326; Duan et al., 1996, 080791; Evans et al., 2003, 

27 048167; Evans et al., 2004, 596379; Fanucchi et al., 2000, 012284; Fanucchi et al., 2006, 096491; 

28 Fujinaka et al., 1985, 040278; Harkema et al., 1987, 041496; Harkema et al., 1987, 040816; 

29 Harkema et al., 1993, 039794; Hatch et al., 1994, 076120; Hyde et al., 1989, 094057; Hyde et al., 

1999, 015124; Joad et al., 2000, 012984; Joad et al., 2006, 596390; Joad et al., 2008, 596391; Larson 

31 et al., 2004, 057062; Lee et al., 1998, 054473; Moffatt et al., 1987, 040841; Plopper and Schelegle, 

32 1997, 656713; Plopper et al., 1991, 042617; Plopper et al., 1998, 087203; Schelegle et al., 2003, 

33 053778; Tran et al., 2004, 628626; Tucker et al., 1993, 056406; Wilson et al., 1984, 040044; Wu et 

34 al., 1999, 012089). 

Since the 1996 and 2006 O3 AQCDs, the initial observations in adult non-human primates 

36 have been expanded in a series of experiments using infant rhesus monkeys repeatedly exposed to 

37	 0.5 ppm O3 starting at 1 month of age (Plopper et al., 2007, 596412). Many of the observations 

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found in adult monkeys have also been noted in infant rhesus monkeys, although a direct comparison 

of the degree of adverse effects between adult and infant monkeys has not been reported. In terms of 

pulmonary function changes, after several episodic exposures of infant monkeys to O3 (each cycle: 

5 days of 0.5 ppm O3 at 8 h/day, followed by 9 days of filtered air exposures), they observed more 

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

airway responsiveness to inhaled histamine (Schelegle et al., 2003, 053778), although neither 

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

inhaled house dust mite antigen alone also produced small but not statistically significant changes in 

baseline airway resistance and airway responsiveness, whereas the combined exposure to both (O3 + 

antigen) produced statistically significant and greater than additive changes in both functional 

measurements. This non-human primate evidence, of an O3-induced change in airway 

responsiveness, supports the biologic plausibility of long_term exposure to O3 contributing to the 

adverse effects of asthma in children. To understand which conducting airways and inflammatory 

mechanisms are involved in O3-induced airway hyperresponsiveness in the infant rhesus monkey, a 

follow-up study examined airway responsiveness ex vivo in lung slices (Joad et al., 2006, 596390). 

Using video microscopy to morphometrically evaluate the response of bronchi and respiratory 

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

responsiveness in asthmatics), the investigators observed differential effects for the two airway sizes. 

While episodic exposure to O3 alone (0.5 ppm) had little effect on ex vivo airway responsiveness in 

bronchi and respiratory bronchioles, exposure to dust mite antigen alone produced airway 

hyperresponsiveness in the large bronchi, whereas O3 + antigen produced significant increases in 

airway hyperresponsiveness only in the respiratory bronchioles. These results suggest that ozone's 

effect on airway responsiveness occurs predominantly in the smaller bronchioles. 

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

O3 alone or O3 + antigen were accompanied by a number of cellular and morphological changes, 

including a significant fourfold increase in eosinophils, (a cell type important in allergic asthma), in 

the bronchoalveolar lavage of infant monkeys exposed to O3 alone. Thus, these studies demonstrate 

both functional and cellular changes in the lung of infant monkeys after cyclic exposure to 0.5 ppm 

O3. This concentration, while higher than those used in controlled human exposure studies, provides 

relevant information to understanding the adverse effects of ambient O3 exposure on the respiratory 

tract of humans. No concentration-response data, however, are available from these non-human 

primate studies. 

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

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

respiratory tract development, conducting airways increase in diameter and length in the infant 

rhesus monkey. Exposure to O3 alone (5 days of 0.5 ppm O3 at 8 h/day, followed by 9 days of 

filtered air exposures for 11 cycles), however, markedly affected the growth pattern of distal 

conducting airways (Fanucchi et al., 2006, 096491). Whereas the first alveolar outpocketing 

occurred at airway generation 13 or 14 in filtered air-control infant monkeys, the most proximal 

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alveolarized airways occurred at an average of 10 airway generations in O3-exposed monkeys. 

Similarly, the diameter and length of the terminal and respiratory bronchioles were significantly 

decreased in O3-exposed monkeys. Importantly, the O3-induced structural pathway changes persisted 

after recovery in filtered air for 6 months after cessation of the O3 exposures. These structural effects 

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

orientation in the respiratory bronchioles, epithelial nerve fiber distribution, and basement membrane 

zone morphometry. These latter effects are significant because of their potential contribution to 

airway obstruction and airway hyperresponsiveness which are central features of asthma. 

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

in infant rhesus monkeys exposed to O3 + house dust mite antigen, but not to O3 alone (Schelegle et 

al., 2003, 053778). To study the underlying mechanisms of this airway hyperresponsiveness, these 

investigators evaluated the effect of exposure to O3 alone and in combination with (+) antigen on 

non-specific airway responsiveness to methacholine at different airway generations. After exposure 

to filtered air, O3, antigen, or O3 + antigen, the bronchi and respiratory bronchioles of 6-month-old 

monkeys were challenged ex vivo with methacholine. Exposure to O3 alone had no significant effect 

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

significant increase in airway responsiveness in the respiratory bronchioles but not the larger 

bronchi. 

Because many cellular and biochemical factors are known to contribute to allergic asthma, the 

effect of exposure to O3 alone or O3 + antigen on immune system parameters was also examined in 

infant rhesus monkeys. Mast cells, which contribute to asthma via the release of potent proteases, 

were elevated in animals exposed to antigen alone but O3 alone had little effect on mast cell numbers 

and the response of animals exposed to O3 + antigen was not different from that of animals exposed 

to antigen alone; thus suggesting that mast cells played little role in the interaction between O3 and 

antigen in this model of allergic asthma (Van Winkle et al., 2010, 670301). Increases in CD4+ and 

CD8+ lymphocytes were observed at 6 months of age in the blood and bronchoalveolar lavage fluid 

of infant rhesus monkeys exposed to O3 + antigen but not in monkeys exposed to either agent alone 

(Miller et al., 2009, 596406). Activated lymphocytes (i.e., CD25+ cells) were morphometrically 

evaluated in the airway mucosa and significantly increased in infant monkeys exposed to antigen 

alone or O3 + antigen. Although O3 alone had no effect on CD25+ cells, it did alter the anatomic 

distribution of CD25+ cells within the airways. Ozone had only a small effect on these sets of 

immune cells and did not produce a strong interaction with an inhaled allergen in this non-human 

primate model; more mechanistic studies are necessary to understand a concentration-response effect 

of O3 on allergic asthma. 

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

cells are thought to play a contributing role to airway hyperresponsiveness. As noted in the 2006 O3 

AQCD, exposure of infant rhesus monkeys altered the normal development of neural innervation in 

the epithelium of the conducting airways (Larson et al., 2004, 057062). Whereas, a significant 

reduction in airway innervation occurred after exposure to O3 alone, a significantly greater reduction 

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1 occurred in monkeys exposed to O3 + antigen. This reduction in overall airway innervation was 

2 accompanied, however, by an increase in the abundance of protein gene product 9.5, a nonspecific 

3 neural marker. Significant increases in protein gene product 9.5 were still observed in O3 alone- and 

4 O3 + antigen-exposed infant monkeys after a 6-month recovery protocol (Kajekar et al., 2007, 

567661). Thus, in addition to structural, immune, and inflammatory effects, exposure to O3 produces 

6 alterations in airway innervation which may contribute to O3-induced exacerbation of asthma. 

7 While the infant rhesus monkey studies examined the effect of long-term O3 exposure on 

8 functional and morphologic development of the lung during early life, a small number of rodent 

9 studies have examined the role of age in the response to O3. In mice, age-related differences in 

O3-induced inflammation and the immediate-early gene response were observed. Johnston and 

11 colleagues (2006, 097439) demonstrated that the lung damage produced by O3 occurred through 

12 distinct (compared to inhaled endotoxin), early gene expression responses. Whereas c-fos and c-jun 

13 mRNA levels were elevated in a concentration-dependent manner (1 and 2.5 ppm O3 for 4 hours) in 

14 the lungs of C57BL/6 mice at 4, 10, and 56 days of age, the relative abundance of mRNA for TLR-4 

(which has been shown to play a role in the pulmonary response to inhaled O3 (Hollingsworth et al., 

16 2010, 635786; Kleeberger et al., 2001, 016163)), was induced in the lungs of 10- and 56-day old but 

17 not 4-day-old mice. Similar age-related differences in response were observed with inhaled 

18 endotoxin, thus suggesting that the murine lung responds differently throughout the postnatal stage 

19 of development. A study by Vancza et al. (2009, 596419) also demonstrated age-related differences 

in the pulmonary response of mice to O3. Significantly greater inflammatory changes were observed 

21 in neonatal (15 to 16 days old) compared to adult (15 week old) mice. Because this increase in 

22 neonatal response was seen only in a subset of the 8 inbred mouse strains exposed to 0.8 ppm O3 for 

23 5 hours, this strain-dependency suggests that genetic host factors play a role in age-related 

24 differences in response to O3. Thus, these rodent studies suggest that the response to O3 in the 

neonatal period is dependent on which postnatal day(s) the exposure occurs, as would be expected in 

26 a rapidly developing mammalian lung. 

27 Collectively, evidence from animal studies strongly suggests that chronic O3 exposure is 

28 capable of damaging the distal airways and proximal alveoli, resulting in lung tissue remodeling ­

29 leading to apparent irreversible changes. Compromised pulmonary function and structural changes 

due to persistent inflammation may exacerbate the progression and development of chronic lung 

31 disease. These findings offer some insight into potential biological mechanisms for the suggested 

32 association between seasonal O3 exposure and reduced lung function development in children as 

33 observed in epidemiologic studies. 

7.2.4. Pulmo na ry In fla mma tion, Inju ry, an d Oxidative Stres s 

34 The 2006 O3 AQCD stated that the extensive human clinical and animal toxicological 

evidence, together with the limited epidemiologic evidence available, suggests a causal role for O3 in 

36 inflammatory responses in the airways. Though the majority of recent studies focus on short-term 

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

of O3-induced inflammation and injury. 

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

children with physician-diagnosed mild persistent asthma in a highly polluted urban area in Italy and 

then again 7 days after relocation to a rural location with significantly lower pollutant levels 

(Renzetti et al., 2009, 199834). The authors observed a fourfold decrease in nasal eosinophils and a 

statistically significant decrease in fractional exhaled nitric oxide along with an improvement in 

lower airway function. Several pollutants were examined, including PM10, NO2, and O3, though 

pollutant-specific results were not presented. These results are consistent with studies showing that 

traffic-related exposures are associated with increased airway inflammation and reduced lung 

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

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

inflammation in large population-based studies (Linn et al., 2009, 597363). Thus, while the time 

scale of 7 days between examinations for eNO needs to be evaluated for appropriateness, the results 

suggest that inflammatory responses are reduced when O3 levels are decreased. 

Chest radiographs (CXR) of 249 children in Mexico City who were chronically exposed to O3 

and PM2.5 were analyzed by Calderón-Garcidueñas et al. (2006, 091253). They reported an 

association between chronic exposures to O3 and other pollutants and a significant increase in 

abnormal CXR’s and lung CTs suggestive of a bronchiolar, peribronchiolar, and/or alveolar duct 

inflammatory process, in clinically healthy children with no risk factors for lung disease. These CXR 

and CT results should be viewed with caution because it is difficult to attribute effects to air 

pollution exposure. 

In a cross-sectional study, Wood et al. (2009, 597085) examined the association of outdoor air 

pollution with respiratory phenotype (PiZZ type) in alpha 1-Antitrypsin deficiency (α-ATD) from the 

U.K. α-ATD registry. In total, 304 PiZZ subjects underwent full lung function testing and 

quantitative high-resolution computed tomography to identify the presence and severity of COPD – 

emphysema. Mean annual air pollution data for 2006 was matched to the location of patients’ houses 

and used in regression models to identify phenotypic associations with pollution controlling for 

covariates. Relative trends in O3 levels were assessed to validate use of a single year’s data to 

indicate long-term exposure and validation; data showed good correlations between modeled and 

measured data (Stedman and Kent, 2008, 110057). Regression models showed that estimated higher 

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

accounting for only a small proportion of the lung function variability. This suggests that a genespecific group demonstrates a long-term O3 exposure effect. 

The similarities of non-human primates to humans make them attractive models in which to 

study the effects of O3 on the respiratory tract. The nasal mucous membranes, which protect the more 

distal regions of the respiratory tract, are susceptible to injury from O3. Carey et al. (2007, 195752) 

conducted a study of O3 exposure in infant rhesus macaques, whose nasal airways closely resemble 

that of humans. Monkeys were exposed either acutely for 5 days (8 h/day) to 0.5 ppm O3, or 

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1 episodically for several biweekly cycles alternating 5 days of 0.5 ppm O3 with 9 days of filtered air 

2 (0 ppm O3), designed to mimic human exposure (70 days total). All monkeys acutely exposed to O3 

3 had moderate to marked necrotizing rhinitis, with focal regions of epitheliar exfoliation, numerous 

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

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

6 exhibited mucous cell metaplasia proximal to the lesions, a protective adaptation observed in adult 

7 monkeys exposed continuously to 0.3 ppm O3 in another study (Harkema et al., 1987, 040816). A 

8 90-day exposure of rats to 0.8 ppm O3 (8 h/day) elicited significantly elevated pro-inflammatory 

9 eicosanoids PGE2 and 12-HETE in BAL, but cytokine profiles did not differ from those of filtered 

air-exposed rats (Schmelzer et al., 2006, 112994). 

7.2.5. Alle rgic Res po n s e s 

11 The association of air pollutants with childhood respiratory allergies was examined in the U.S. 

12 using the 1999-2005 National Health Interview Survey of approximately 70,000 children, and 

13 ambient air pollution data from the U.S. EPA, with monitors within 20 miles of each child’s 

14 residential block (Parker et al., 2009, 192359). The authors examined the associations between the 

reporting of respiratory allergy or hay fever and summer exposure to O3, controlling for 

16 demographic and geographic factors. Increased respiratory allergy/hay fever was associated with 

17 increased O3 levels (adjusted OR per 10 ppb = 1.20; [95% CI: 1.15, 1.26]). These associations 

18 persisted after stratification by urban-rural status, inclusion of multiple pollutants, and definition of 

19 exposure by differing exposure radii; smaller samples within 5 miles of monitors were remarkably 

similar to the primary results. No associations between the other pollutants and the reporting of 

21 respiratory allergy/hay fever were apparent. Ramadour et al. (2000, 013259) reported no relationship 

22 between O3 levels and rhinitis symptoms and hay fever. Hwang et al. (2006, 088971) report the 

23 prevalence of allergic rhinitis (adjusted OR per 10 ppb = 1.05; [95% CI: 0.98, 1.12]) in a large cross24 sectional study in Taiwan. In a large cross-sectional study in France, Penard-Morand et al. (2005, 

087951) reported a positive relationship between lifetime allergic rhinitis and O3 exposure in a two26 pollutant model with NO2. These studies related positive outcomes of allergic response and O3 

27 exposure but with variable strength for the effect estimates. Nasal eosinophils, which participate in 

28 allergic disease, were observed to decrease by fourfold in 37 atopic, mildly asthmatic children 7 days 

29 after relocation from a highly polluted urban area in Italy to a rural location with significantly lower 

pollutant levels (Renzetti et al., 2009, 199834). 

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

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

33 (Rage et al., 2009, 196719). Geostatistical models were performed on 4×4 km grids to assess 

34 individual outdoor air pollution exposure that was assigned to subject’s home address. Ozone 

concentrations were positively related to total IgE levels and an increase of 5 ppb of O3 resulted in an 

36 increase of 20.4% (95% CI: 3.0, 40.7) in total IgE levels. Nearly 75% of the subjects were atopic. 

37 Two-pollutant models for O3 with NO2 were decreased by 25% while NO2 was decreased by 57%. 

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1 Associations were not sensitive to adjustment for covariates or the season of IgE measurements. 

2 These cross-sectional results suggest that exposure to O3 may increase total IgE in adult asthmatics. 

3 No toxicological studies of long-term exposure are available, but short-term exposure studies 

4 in rodents and non-human primates demonstrate allergic skewing of immune responses and 

enhanced IgE production. Due to the persistent nature of these responses, the short-term 

6 toxicological evidence lends biological plausibility to the limited epidemiologic findings of an 

7 association between long-term O3 exposure and allergic outcomes. 

7.2.6. Ho s t Defe ns e 

8 Short-term exposures to O3 cause decreases in host defenses against infectious lung disease in 

9 animal models. However, acute O3-induced suppression of alveolar phagocytosis and immune 

functions observed in animals appears to be transient and attenuated with continuous or repeated 

11 exposures. Chronic exposures (weeks, months) of 0.1 ppm do not cause greater effects on infectivity 

12 than short exposures, due to defense parameters becoming reestablished with prolonged exposures, 

13 although chronic exposure has been shown to slow alveolar clearance. In an older study (Jakab and 

14 Bassett, 1990, 042196), no detrimental effects were seen with a 120-day exposure to 0.5 ppm O3 on 

acute lung injury from influenza virus administered immediately before O3 exposure started. But 

16 there were O3-enhanced postinfluenzal alveolitis and lung parenchymal changes. No new evidence 

17 has become available to address the effects of long-term exposure on host defense mechanisms. 

7.2.7. Re s pirato ry Mo rtality 

18 A limited number of epidemiologic studies have assessed the relationship between long-term 

19 exposure to O3 and mortality. The 2006 O3 AQCD concluded that an insufficient amount of evidence 

existed “to suggest a causal relationship between chronic O3 exposure and increased risk for 

21 mortality in humans” (U.S. EPA, 2006, 088089). Though total and cardio-pulmonary mortality were 

22 considered in these studies, respiratory mortality was not specifically considered. In the most recent 

23 follow-up analysis of the ACS cohort (Jerrett et al., 2009, 194160), cardiopulmonary deaths were 

24 subdivided into respiratory and cardiovascular, separately, as opposed to combined in the Pope et al. 

(2002, 024689) work. A 10-ppb increment in exposure to O3 elevated the risk of death from 

26 respiratory causes and this effect was robust to the inclusion of PM2.5. The association between 

27 increased O3 concentrations and increased risk of death from respiratory causes was insensitive to 

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

29 area and state of residence, and to adjustment for several ecologic variables considered individually. 

7.2.8. Su mma ry an d Cau s al De te rmina tio n 

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

31 long-term O3 exposures and asthma-related symptoms, asthma prevalence, or allergy to common 

32 aeroallergens among children after controlling for covariates. Little evidence was available to relate 

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1 long-term exposure to current ambient O3 concentrations to deficits in the growth rate of lung 

2 function in children. Additionally, limited evidence was available evaluating the relationship 

3 between long-term O3 levels and pulmonary inflammation and other endpoints. From toxicological 

4 studies, it appeared that O3-induced inflammation tapered off during long-term exposures, but that 

5 hyperplastic and fibrotic changes remained elevated and in some cases even worsened after a 

6 postexposure period in clean air. Episodic exposures were also known to cause more severe 

7 pulmonary morphologic changes than continuous exposure (U.S. EPA, 2006, 088089). 

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

9 analysis methods evaluating the relationship between long-term measures of exposure to ambient O3 

10 and measures of respiratory morbidity conducted by different research groups in different locations. 

11 See Table 7-1 for O3 concentrations associated with selected studies. The positive results from 

12 various designs and locations support an association between long-term O3 concentrations and 

13 respiratory morbidity. 

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

15 in children. Studies have provided evidence for a relationship between different genetic variants 

16 (HMOX, GST, ARG) that, in combination with O3 exposure, are related to new onset asthma (Islam 

17 et al., 2008, 097348; Islam et al., 2009, 196715; Salam et al., 2009, 596644). These studies involve 

18 two separate cohorts in the CHS. These prospective cohort studies represent strong evidence because 

19 they are methodologically rigorous epidemiology studies. The studies were conducted in 12 

20 California communities. The stratified analysis for the two independent fourth-grade cohorts of the 

21 study population recruited in 1993 and 1996 yielded consistent results and provides replication in 

22 independent groups of children. Also, no meaningful interactions were observed between other air 

23 pollutants such as PM10 and genes. 

Table 7-1. Summary of selected key new studies examining annual ozone exposure and respiratory 

health effects 

Study; Health Effect; Location Mean Annual O3 Concentration (ppb) O3 Range (ppb) 

Percentiles 

Islam et al.(2008, 097348); 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, 196715); 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, 596644); childhood onset asthma; CHS O3 greater than or less than 50 ppb See left 

Lin et al. (2008, 196680); 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, 196685); asthma 

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

Meng et al. (2010, 594252); 

asthma ED visits or hospitalizations; 

San Joaquin Valley, CA 

Median 30.3 ppb 25-75% range 

27.1 to 34.0 

Lee et al. (2009, 199915); 

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

Rage et al. (2009, 196719); 

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

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Wenten et al. (2009, 597084); Min-Max Median 46.9 ppb; 10:00 a.m. – 6:00 p.m. respiratory school absence, US 27.6-65.3 

Studies using a cross-sectional design provide support for a relationship between long-term O3 

exposure and health effects in asthmatics. A long-term O3 exposure study relates bronchitic 

symptoms to TNF-308 genotype asthmatic children with ambient O3 exposure in the CHS (Lee et al., 

2009, 199915). A study relating asthma severity to long-term O3 exposure in five French cities 

provides additional support to the notion that effects on asthma are related to long-term O3 exposure 

(Rage et al., 2009, 196720). For the respiratory health of the general U.S. population, risk of 

respiratory-related school absences was elevated for children with the CAT and MPO variant genes 

related to communities with high ambient O3 levels (Wenten et al., 2009, 597084). 

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

concentration-response relationship in a New Yor k State birth cohort (Lin et al., 2008, 196680). A 

separate hospitalization cross-sectional study in San Joaquin Valley in California reports similar 

findings (Meng et al., 2010, 594252). Another study relates asthma hospital admissions to quarterly 

average O3 in the South Coast Air Basin of California (Moore et al., 2008, 196685). 

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

or development can result in irreversible morphological changes in the lung, which in turn can 

influence pulmonary function. Studies by Plopper and colleagues have demonstrated changes in 

pulmonary function and airway morphology in adult and infant non-human primates repeatedly 

exposed to environmentally relevant concentrations of O3 (Fanucchi et al., 2006, 096491; Harkema 

et al., 1987, 041496; Joad et al., 2006, 596390; Schelegle et al., 2003, 053778). This non-human 

primate evidence of an O3-induced change in airway responsiveness supports the biologic 

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

Results from epidemiologic studies examining long-term O3 exposure and pulmonary function 

effects are inconclusive with some new studies relating effects at higher exposure levels. The results 

from the CHS still remain as the definitive line of evidence. Other cross-sectional studies provide 

mixed results. 

The 2006 O3 AQCD states that the extensive human clinical and animal toxicological 

evidence, together with the limited epidemiologic evidence available, suggests a causal role for O3 in 

inflammatory responses in the airways. Though the majority of recent studies focus on short-term 

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

of O3-induced inflammation and injury. Toxicological studies in rodents and non-human primates 

indicate that chronic O3 exposure causes structural changes in the respiratory tract, and simulated 

seasonal exposure studies suggest that such exposures might have cumulative impacts. The strongest 

epidemiologic evidence for a relationship between long-term O3 exposure and respiratory morbidity 

is provided by new studies that demonstrate associations between long-term measures of O3 

exposure and new-onset asthma in children and increased respiratory symptom effects in asthmatics. 

While there are currently a limited number of studies in this data base, these U.S. multi-community 

prospective cohort studies are methodologically rigorous epidemiology studies. Asthma risk is 

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1 related to the important relationships between genetic variability, environmental O3 exposure, and 

2 behavior. These relationships are complex. The genes, evaluated in these studies, are both key 

3 candidates in the oxidative stress pathway and have been shown to play an important role in asthma 

4 development. Reduced risk for asthma development is reported in some studies in children living in 

5 low- O3 communities. Ozone levels in the studies (10:00 a.m. to 6:00 p.m.) ranged from 28.6 to 

6 45.5 ppb in low O3 communities (mean = 38.4 ppb) and from 46.5 to 64.9 ppb in high O3 

7 communities (mean = 55.2 ppb). Other studies in the new data base provide coherent evidence for 

8 long-term O3 exposure and respiratory morbidity effects such as first asthma hospitalization and 

9 respiratory symptoms in asthmatics. Studies considering other pollutants provide data suggesting 

10 that the effects related to O3 are independent from potential effects of the other pollutants. Some 

11 studies provide evidence for a positive concentration-response relationship. The above discussion of 

12 the recent epidemiologic and toxicological data base provides a compelling case to support the 

13 hypothesis that a relationship exists between long-term exposure to ambient O3 and measures of 

14 respiratory morbidity. The 2006 O3 AQCD concluded the evidence was suggestive but inconclusive 

15 at that time. The new epidemiological data base, combined with toxicological studies in rodents and 

16 non-human primates, provides biologically plausible evidence that there is likely to be causal 

17 relationship between long-term exposure to O3 and respiratory morbidity. 

7.3. Cardiovas cular Effects 

7.3.1. Ca rdio va s cular Dis ea s e 

7.3.1.1. Cardiovas cular Epide miology 

18 Long-term exposure to O3 and its effects on cardiovascular morbidity were not considered in 

19 the 2006 O3 AQCD (U.S. EPA, 2006, 088089). However, recent studies have assessed the chronic 

20 effects of O3 exposure on cardiovascular morbidity (Chen et al., 2007, 145956; Chuang et al., 2011, 

21 670846; Forbes et al., 2009, 190351). The association between O3 exposure and markers of lipid 

22 peroxidation and antioxidant capacity was examined among 120 nonsmoking healthy college 

23 students, aged 18-22 years, from the University of California, Berkeley (Feb-Jun 2002) (Chen et al., 

24 2007, 145956). By design, students were chosen that had experienced different geographic levels of 

25 O3 over their lifetimes and during recent summer vacation in either greater Los Angeles (LA) or the 

26 San Francisco Bay Area (SF). A marker of lipid peroxidation, 8-isoprostane (8-iso-PGF) in plasma, 

27 was assessed. This marker is formed continuously under normal physiological conditions but has 

28 been found at elevated concentrations in response to environmental exposures. A marker of overall 

29 antioxidant capacity, ferric reducing ability of plasma (FRAP), was also measured. The lifetime O3 

30 exposure estimates (estimated monthly average, ppb) did not show much overlap between the two 

31 geographic areas [median (range): LA, 42.9 (28.5-65.3); SF, 26.9 (17.6-33.5)]. Estimated lifetime O3 

32 exposure was related to 8-iso-PGF [β = 0.025 (pg/mL)/8-h ppb O3, p = 0.0007]. For the 17-ppb 

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1 cumulative lifetime O3 exposure difference between LA and SF participants, there was a 

2 17.41-pg/mL (95% CI: 15.43, 19.39) increase in 8-iso-PGF. No evidence of association was 

3 observed between lifetime O3 exposure and FRAP [β = -2.21 (pg/mL)/8-h ppb O3, p = 0.45]. The 

4 authors note that O3 was highly correlated with PM10-2.5 and NO2 in this study population; however, 

their inclusion in the O3 models did not substantially modify the magnitude of the associations with 

6 O3. Because the lifetime exposure results were supported by shorter-term exposure results from 

7 analyses considering O3 concentrations up to 30 days prior to sampling, the authors conclude that 

8 persistent exposure to O3 can lead to sustained oxidative stress and increased lipid peroxidation. 

9 However, because there was not much overlap in lifetime O3 exposure estimates between LA and SF, 

it is possible that the risk estimates involving the lifetime O3 exposures could be confounded by 

11 unmeasured factors related to other differences between the two cities. 

12 Forbes et al. (2009, 190351) used the annual average exposures to assess the relationship 

13 between chronic ambient air pollution and levels of fibrinogen and C-reactive protein (CRP) in a 

14 cross-sectional study conducted in England. Data were collected from the Health Survey of England 

for 1994, 1998, and 2003. The sampling strategy was designed to obtain a representative sample of 

16 the English population; however, due to small group sizes, only data from white ethnic groups were 

17 analyzed. For analyses, the annual concentrations of O3 were averaged for the year of data collection 

18 and the previous year with the exception of 1994 (because pollutant data were not available for 

19 1993). Median O3 concentrations were 26.7 ppb, 25.4 ppb, and 28 ppb for 1994, 1998, and 2003, 

respectively. Year specific adjusted effect estimates were created and combined in a meta-analysis. 

21 No evidence of association was observed for O3 and levels of fibrinogen or CRP (e.g., the combined 

22 estimates for the percent change in fibrinogen and CRP for a 10 ppb increase in O3 were -0.28 [95% 

23 CI: -2.43, 1.92] and -3.05 [95% CI: -16.10, 12.02], respectively). Further research will be important 

24 for understanding the effects, if any, of chronic O3 exposure on cardiovascular morbidity risk. 

A study was performed in Taiwan to examine the association between long-term O3 

26 concentrations and blood pressure and blood markers using the Social Environment and Biomarkers 

27 of Aging Study (SEBAS) (Chuang et al., 2011, 670846). Individuals included in the study were 

28 54 years of age and older. The mean annual O3 concentration during the study period was 22.95 ppb 

29 (SD 6.76 ppb). Positive associations were observed between O3 concentrations and both systolic and 

diastolic blood pressure [changes in systolic and diastolic blood pressure were 21.51mmHg (95% CI: 

31 16.90, 26.13) and 20.56 mmHg (95% CI: 18.14, 22.97) per 8.95 ppb increase in O3, respectively). 

32 Increased O3 concentrations were also associated with increased levels of total cholesterol, fasting 

33 glucose, hemoglobin A1c, and neutrophils. No associations were observed between O3 

34 concentrations and triglyceride and IL-6 levels. The observed associations were reduced when other 

pollutants were added to the models. 

7.3.1.2. Long-Term Ca rdiovas cular Toxicology 

36 Three new studies have investigated the cardiovascular effects of long-term exposure to O3 in 

37 animal models. In addition to the short-term effects described in Section 6.3.1, a recent study found 

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that ApoE-/- mice (6-14 weeks old) exposed to O3 (0.5 ppm) for 8 weeks (5 days/week, 8 h/day) had 

enhanced aortic atherosclerotic lesion area compared to air exposed controls (Chuang et al., 2009, 

197202). Chuang et al. (2009, 197202) not only provided evidence for increased atherogenesis in 

susceptible mice, but also reported an elevated vascular inflammatory and redox state in wild-type 

mice and infant primates (Section 6.3.1). This study is compelling in that it identifies biochemical 

and cellular events responsible for transducing the airway epithelial reactions of O3 into 

proinflammatory responses that are apparent in the extrapulmonary vasculature (Cole and Freeman, 

2009, 597507). 

Another recent study provides further evidence for increased vascular inflammation and 

oxidation and long term effects in the extrapulmonary space. Rats episodically exposed to O3 

(0.4 ppm) for 16 weeks (5 h/day, 1 day/week) presented marked increases in gene expression of 

biomarkers of oxidative stress, thrombosis, vasoconstriction, and proteolysis (Kodavanti et al., In 

Press, 666323). Ozone exposure upregulated aortic mRNA expression of heme oxygenase-1 (HO-1), 

tissue plasminogen activator (tPA), plasminogen activator inhibitor-1 (PAI-1), von Willebrand factor 

(vWf), thrombomodulin, endothelial nitric oxide synthase (eNOS), endothelin-1 (ET-1), matrix 

metalloprotease-2 (MMP-2), matrix metalloprotease-3 (MMP-3), and tissue inhibitor of matrix 

metalloprotease-2 (TIMP-2). In addition, O3 exposure depleted some cardiac mitochondrial 

phospholipid fatty acids (C16:0 and C18:1), which may be the result of oxidative modifications. The 

authors speculate that oxidatively modified lipids and proteins produced in the lung and heart 

promote vascular pathology through activation of lectin-like oxidized-low density lipoprotein 

receptor-1 (LOX-1). Activated LOX-1 induces expression of a number of the biomarkers induced by 

O3 exposure and is considered pro-atherogenic. Both LOX-1 mRNA and protein were increased in 

mouse aorta after O3 exposure. This study provides a possible pathway and further support to the 

observed O3 induced atherosclerosis. 

Vascular occlusion resulting from atherosclerosis can block blood flow through vessels 

causing ischemia. The restoration of blood flow or reperfusion can cause injury to the tissue from 

subsequent inflammation and oxidative damage. Ozone exposure (0.8 ppm for 28 or 56 days) 

enhanced the sensitivity to myocardial ischemia-reperfusion (I/R) injury in Sprague-Dawley rats 

while increasing oxidative stress levels and pro-inflammatory mediators and decreasing production 

of anti-inflammatory proteins (Perepu et al., 2010, 385020). Both long- and short-term O3 exposure 

decreased the left ventricular developed pressure, rate of change of pressure development, and rate 

of change of pressure decay and increased left ventricular end diastolic pressure in isolated perfused 

hearts. In this ex vivo heart model, O3 induced oxidative stress by decreasing SOD enzyme activity 

and increasing malondialdehyde levels. Ozone also elicited a proinflammatory state evident by an 

increase in TNF-α and a decrease in the anti-inflammatory cytokine IL-10. The authors conclude that 

O3 exposure will result in a greater I/R injury. 

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7.3.2. Ca rdia c Mo rtality 

1 A limited number of epidemiologic studies have assessed the relationship between long-term 

2 exposure to O3 and mortality. The 2006 O3 AQCD concluded that an insufficient amount of evidence 

3 existed “to suggest a causal relationship between chronic O3 exposure and increased risk for 

4 mortality in humans” (U.S. EPA, 2006, 088089). Though total and cardio-pulmonary mortality were 

considered in these studies, cardiovascular mortality was not specifically considered. In the most 

6 recent follow-up analysis of the ACS cohort (Jerrett et al., 2009, 194160), cardiopulmonary deaths 

7 were subdivided into respiratory and cardiovascular, separately, as opposed to combined in the Pope 

8 et al. (2002, 024689) work. A 10-ppb increment in exposure to O3 elevated the risk of death from the 

9 cardiopulmonary, cardiovascular, and ischemic heart disease. Inclusion of PM2.5 as a co-pollutant 

attenuated the association with exposure to O3 for all of the cardiovascular endpoints to become null. 

7.3.3. Summary and Causal De te rmina tio n 

11 Previous AQCDs did not address the cardiovascular effects of long-term O3 exposure due to 

12 limited data availability. The evidence remains limited; however the emerging data is supportive of a 

13 role for O3 in chronic cardiovascular diseases. Few epidemiologic studies have investigated 

14 cardiovascular morbidity after long-term O3 exposure, and the majority only assessed cardiovascular 

disease related biomarkers. A study on O3 and cardiovascular mortality reported no association after 

16 adjustment for PM2.5 levels. Further epidemiologic studies on cardiovascular morbidity and mortality 

17 after long-term exposure have not been published. 

18 Toxicological evidence on long-term O3 exposure is also limited but three strong toxicological 

19 studies have been published since the previous AQCD. These studies provide evidence for O3 

enhanced atherosclerosis and I/R injury, corresponding with development of a systemic oxidative, 

21 proinflammatory environment. Although questions exist for how O3 inhalation causes systemic 

22 effects, a recent study proposes a mechanism for development of vascular pathology that involves 

23 activation of LOX-1 by O3 oxidized lipids and proteins. This activation may also be responsible for 

24 O3 induced changes in genes involved in proteolysis, thrombosis, and vasoconstriction. Taking into 

consideration the positive toxicological studies reported, the generally limited body of evidence is 

26 suggestive of a causal relationship between relevant long-term exposures to O3 and 

27 cardiovascular effects. 

7.4. Reproductive and Developmental Effects 

28 Although the body of literature is growing, the research focusing on adverse birth outcomes is 

29 limited when compared to the numerous studies that have examined the more well-established health 

effects of air pollution. Among this small number of studies, various measures of birth weight and 

31 fetal growth, such as low birth weight (LBW), small for gestational age (SGA), and intrauterine 

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1 growth restriction (IUGR), and preterm birth (<37-week gestation; [PTB]) have received more 

2 attention in air pollution research, while congenital malformations are less studied. 

3 Infants and fetal development processes may be particularly vulnerable to O3 exposure, and 

4 although the physical mechanisms are not fully understood, several hypotheses have been proposed 

involving direct effects on fetal health, altered placenta function, or indirect effects on the mother’s 

6 health (Bracken et al., 2003, 156288; Clifton et al., 2001, 156360; Maisonet et al., 2004, 156725; 

7 Schatz et al., 1990, 156073; Sram et al., 2005, 087442). Study of these outcomes can be difficult 

8 given the need for detailed exposure data and potential residential movement of mothers during 

9 pregnancy. Air pollution epidemiologic studies reviewed in the 2006 O3 AQCD examined impacts on 

birth-related endpoints, including intrauterine, perinatal, postneonatal, and infant deaths; premature 

11 births; intrauterine growth retardation; very low birth weight (weight <1,500 grams) and low birth 

12 weight (weight <2,500 grams); and birth defects. However, in the limited number of studies that 

13 investigated O3, no associations were found between O3 and birth outcomes, with the possible 

14 exception of birth defects. 

Two recent articles have reviewed methodological issues relating to the study of outdoor air 

16 pollution and adverse birth outcomes (Ritz and Wilhelm, 2008, 156914; Slama et al., 2008, 156985). 

17 Some of the key challenges to interpretation of these study results include the difficulty in assessing 

18 exposure as most studies use existing monitoring networks to estimate individual exposure to 

19 ambient air pollution; the inability to control for potential confounders such as other risk factors that 

affect birth outcomes (e.g., smoking); evaluating the exposure window (e.g., trimester) of 

21 importance; and limited evidence on the physiological mechanism of these effects (Ritz and 

22 Wilhelm, 2008, 156914; Slama et al., 2008, 156985). Although early animal studies (Kavlock et al., 

23 1980, 094043) found that exposure to O3 in the late gestation of pregnancy in rats led to some 

24 abnormal reproductive performances for neonates, to date human studies have reported inconsistent 

results for the association of ambient O3 on birth outcomes. 

7.4.1. Effects on Sperm 

26 A limited amount of research has been conducted to examine the association between air 

27 pollution and male reproductive outcomes, specifically semen quality. To date, the epidemiologic 

28 studies have considered various exposure durations before semen collection that encompass either 

29 the entire period of spermatogenesis (i.e., 90 days) or key periods of sperm development that 

correspond to epididymal storage, development of sperm motility, and spermatogenesis. In an 

31 analysis conducted as part of the Teplice Program, 18-year-old men residing in the heavily polluted 

32 district of Teplice in the Czech Republic were found to be at greater risk of having abnormalities in 

33 sperm morphology and chromatin integrity than men of similar age residing in Prachatice, a less 

34 polluted district (Selevan et al., 2000, 012578; Sram et al., 1999, 078127). A follow-up longitudinal 

study conducted on a subset of the same men from Teplice revealed associations between total 

36 episodic air pollution and abnormalities in sperm chromatin (Rubes et al., 2005, 078091). A 

37 limitation of these studies is that they did not identify specific pollutants and their concentrations. 

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1 More recent epidemiologic studies conducted in the U.S. have also reported associations 

2 between ambient air pollution and sperm quality for individual air pollutants, including O3 and 

3 PM2.5. In a repeated measures study in Los Angeles, CA, Sokol et al. (2006, 098539) reported a 

4 reduction in average sperm concentration during three exposure windows (0-9, 10-14, and 

70-90 days before semen collection) associated with high ambient levels of O3 in healthy sperm 

6 donors. This effect persisted under a joint additive model for O3, CO, NO2 and PM10. The authors did 

7 not detect a reduction in sperm count. Hansen et al. (2010, 594438) investigated the effect of 

8 exposure to O3 and PM2.5 on sperm quality in three southeastern counties (Wake County, NC; Shelby 

9 County, TN; Galveston County, TX). Outcomes included sperm concentration and count, 

morphology, DNA integrity and chromatin maturity. Overall, the authors found both protective and 

11 adverse effects, although some results suggested adverse effects on sperm concentration, count and 

12 morphology. There was evidence of an association between decreased sperm concentration and 

13 count with O3, though these associations were not statistically significant. 

14 The biological mechanisms linking ambient air pollution to decreased sperm quality have yet 

to be determined, though O3-induced oxidative stress, inflammatory reactions, and the induction of 

16 the formation of circulating toxic species have been suggested as possible mechanisms (Sokol et al., 

17 2006, 098539). Decremental effects on testicular morphology have been demonstrated in 

18 toxicological studies with histological evidence of O3-induced depletion of germ cells in testicular 

19 tissue and decreased seminiferous tubule epithelial layer. Jedlinska-Krakowska et al. (2006, 195640) 

demonstrated histopathological evidence of impaired spermatogenesis (round spermatids/ 

21 spermatocytes, giant spermatid cells, and focal epithelial desquamation with denudation to the 

22 basement membrane). The exposure protocol used five month old adult rats exposed to O3 as adults 

23 (0.5 ppm, 5 h/day for 50 days). This degeneration could be rescued by vitamin E administration, 

24 indicating an antioxidant effect. Vitamin C administration had no effect at low doses of ascorbic acid 

and exacerbated the O3-dependent damage at high doses, as would be expected as vitamin C can be a 

26 radical generator instead of an antioxidant at higher doses. In summary, this study provided 

27 toxicological evidence of impaired spermatogenesis with O3 exposure that was rescued with certain 

28 antioxidant supplementation. 

29 Overall, there is limited epidemiologic evidence for an association with O3 concentration and 

decreased sperm concentration. A recent toxicological study provides limited evidence for a possible 

31 biological mechanism (histopathology showing impaired spermatogenesis) for such an association. 

7.4.2. Effects on Reproduction 

32 Evidence suggests that exposure to air pollutants during pregnancy is associated with adverse 

33 birth outcomes, which has been attributed to the increased susceptibility of the fetus due to 

34 physiologic immaturity. Gametes (i.e., ova and sperm) may be even more vulnerable, especially 

outside of the human body, as occurs with assisted reproduction. Smokers require twice the number 

36 of in vitro fertilization (IVF) attempts to conceive as non-smokers (Feichtinger et al., 1997, 625251), 

37 suggesting that a preconception exposure can be harmful to pregnancy. A recent study used an 

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1 established national-scale, log-normal kriging method to spatially estimate daily mean 

2 concentrations of criteria pollutants at addresses of women undergoing their first IVF cycle and at 

3 their IVF labs from 2000 to 2007 in the northeastern U.S. (Legro et al., 2010, 597377). Increasing O3 

4 concentration at the patient's address was significantly associated with an increased chance of live 

birth during ovulation induction (OR 1.13, [95% CI: 1.05, 1.22] per 10 ppb increase), but with 

6 decreased odds of live birth when exposed from embryo transfer to live birth (OR 0.79, [95% CI: 

7 0.69, 0.90] per 10 ppb increase). After controlling for NO2 in a co-pollutant model, however, O3 was 

8 no longer significantly associated with IVF failure. The results of this study suggest that exposure to 

9 O3 during ovulation was beneficial (perhaps due to early conditioning to O3), whereas later exposure 

to O3 (e.g., during gestation) was detrimental, and reduced the likelihood of a live birth. 

11 In toxicological studies, reproductive success in rats appears unaffected by O3 exposure. 

12 Ozone administration (continuous 0.4, 0.8 or 1.2 ppm O3) to CD-1 mouse dams during the majority 

13 of pregnancy (PD7-17, which excludes the pre-implantation period), led to no adverse effects on 

14 reproductive success (proportion of successful pregnancies, litter size, sex ratio, frequency of still 

birth, or neonatal mortality) (Bignami et al., 1994, 076063). There was a nearly statistically 

16 significant increase in pregnancy duration (0.8 and 1.2 ppm O3). Initially, dam body weight (0.8 and 

17 1.2 ppm), water consumption (0.4, 0.8 and 1.2 ppm O3) and food consumption (0.4, 0.8 and 1.2 ppm) 

18 during pregnancy were decreased with O3 exposure but these deficits dissipated a week or two after 

19 the initial exposure (Bignami et al., 1994, 076063). The anorexigenic effect of O3 exposure on the 

pregnant dam appears to dissipate with time; the dams seem to adapt to the O3 exposure. In males, 

21 data exist showing morphological evidence of altered spermatogenesis in O3 exposed animals 

22 (Jedlinska-Krakowska et al. (2006, 195640). Some evidence suggests that O3 may affect 

23 reproductive success when combined with other chemicals. Kavlock et al. (1979, 039228) showed 

24 that O3 acted synergistically with sodium salicylate to increase the rate of pup resorptions after 

midgestational exposure (1.0 ppm O3, GD9-12). At low doses of O3 exposure, toxicological studies 

26 show reproductive effects to include a transient anorexigenic effect of O3 on gestational weight gain, 

27 and a synergistic effect of O3 on salicylate-induced pup resorptions; other fecundity, pregnancy and 

28 gestation related outcomes appear unaffected by O3 exposure. Collectively, there is very little 

29 epidemiologic evidence for the effect of O3 on reproductive success, and the reproductive success in 

rats appears to be unaffected in toxicological studies of O3 exposure. 

7.4.3. Birth Weigh t 

31 With birth weight routinely collected in vital statistics and being a powerful predictor of infant 

32 mortality, it is the most studied outcome within air pollution-birth outcome research. Air pollution 

33 researchers have analyzed birth weight as a continuous variable and/or as a dichotomized variable in 

34 the form of LBW (<2,500 g [5 lbs, 8 oz]). 

Birth weight is primarily determined by gestational age and intrauterine growth, but also 

36 depends on maternal, placental and fetal factors as well as on environmental influences. In both 

37 developed and developing countries, LBW is the most important predictor for neonatal mortality and 

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1 is a significant determinant of postneonatal mortality and morbidity. Recent studies report that 

2 infants who are smallest at birth have a higher incidence of diseases and disabilities, which continue 

3 into adulthood (Hack and Fanaroff, 1999, 625952). 

4 The strongest evidence for an effect of O3 on birth weight comes from the Children’s Health 

5 Study conducted in southern California. In this study, Salam et al. (2005, 087885) report that 

6 maternal exposure to O3 averaged over the entire pregnancy was associated with reduced birth 

7 weight for 24-h avg (39.3 g decrease [95% CI: -55.8, -22.8] in birth weight per 10 ppb and 8-h avg 

8 (19.2-g decrease [95% CI: -27.7, -10.7] in birth weight per 10 ppb) O3 concentrations. This effect 

9 was stronger for concentrations averaged over the second and third trimesters. PM10, NO2 and CO 

10 concentrations averaged over the entire pregnancy were not statistically significantly associated with 

11 birth weight, though CO concentrations in the first trimester and PM10 concentrations in the third 

12 trimester were associated with a decrease in birth weight. Additionally, the authors observed a 

13 concentration-response relationship of birth weight with 24-h avg O3 concentrations averaged over 

14 the entire pregnancy that was clearest above the 30-ppb level (see Figure 7-4). Relative to the lowest 

15 decile of 24-h avg O3, estimates for the next 5 lowest deciles were approximately -40 g to -50 g, with 

16 no clear trend and with 95% confidence bounds that included zero. The highest four deciles of O3 

17 exposure showed an approximately linear decrease in birth weight, and all four 95% CIs excluded 

18 zero, and ranged from mean decreases of 74 grams to decreases of 148 grams. 

Source: Salam et al. (2005, 087885) 

Figure 7-4. Birthweight deficit by decile of 24-h avg ozone concentration averaged over the

entire pregnancy compared with the decile group with the lowest ozone exposure. ﷒

Deficits are plotted against the decile-group-specific median ozone exposure. Error 

bars represent 95% CIs. Indicator variables for each decile of ozone exposure

(except the least-exposed group) were included in a mixed model. 

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Several additional studies conducted in the U.S. and Canada also investigated the association 

between ambient O3 concentrations and birth weight and found little evidence for an association. 

Morello-Frosch et al. (2010, 670076) estimated ambient O3 concentrations throughout pregnancy 

and for each trimester in the neighborhoods of women who delivered term singleton births between 

1996 and 2006 in California. A 10-ppb increase in O3 averaged across the entire pregnancy was 

associated with a 5.7-g decrease (95% CI: -6.6, -4.9) in birth weight when exposures were calculated 

using monitors within 10 km of the maternal address at date of birth. When the distance from the 

monitor was restricted to 3 km, the decrease in birth weight associated with a 10-ppb increase in O3 

increased to 8.9 g (95% CI: -10.6, -7.1). These results persisted in co-pollutant models and in models 

that stratified by trimester of exposure, SES, and race. Chen et al. (2002, 024945) used 8-h avg O3 

concentrations to create exposure variables based on average maternal exposure for each trimester. 

Ozone was not found to be related to birth weight in single-pollutant models, though the O3 effect 

during the third trimester was borderline significant in a co-pollutant model with PM10. Wilhelm and 

Ritz (2005, 088668) extended previous analyses of term LBW (Ritz and Yu, 1999, 086976; Ritz et 

al., 2000, 012068) to include the period 1994-2000. The authors examined varying residential 

distances from monitoring stations to see if the distance affected risk estimation, exploring the 

possibility that effect attenuation may result from local pollutant heterogeneity inadequately captured 

by ambient monitors. As in their previous studies, the authors observed associations between 

elevated concentrations of CO and PM10 both early and late in pregnancy and risk of term LBW. 

After adjusting for CO and/or PM10 the authors did not observe associations between O3 and term 

LBW in any of their models. Brauer et al. (2008, 156292) evaluated the impacts of air pollution (CO, 

NO2, NO, O3, SO2, PM2.5, PM10) on birth weight for the period 1999-2002 using spatiotemporal 

residential exposure metrics by month of pregnancy in Vancouver, BC. Quantitative results were not 

presented for the association between O3 and LBW, though the authors observed associations that 

were largely protective. Dugandzic et al. (2006, 088681) examined the association between LBW 

and ambient levels of air pollutants by trimester of exposure among a cohort of term singleton births 

from 1988-2000. Though there was some indication of an association with SO2 and PM10, there were 

no effects for O3. 

Similarly, studies conducted in Australia, Latin America, and Asia report limited evidence for 

an association between ambient O3 and measures of birth weight. In Sydney, Australia, Mannes et al. 

(2005, 087895) found that O3 concentrations in the second trimester of pregnancy had small adverse 

effects on birth weight (7.5-g decrease; [95 % CI: -13.8, 1.2] per 10 ppb), though this effect 

disappeared when the analysis was limited to births with a maternal address within 5 km of a 

monitoring station (87.7-g increase; [95% CI: 10.5, 164.9] per 10 ppb). Hansen et al. (2007, 090703) 

reported that trimester and monthly specific exposures to all pollutants were not statistically 

significantly associated with a reduction in birth weight in Brisbane, Australia. In Sao Paulo, Brazil, 

Gouveia et al. (2004, 055613) found that O3 exhibited a small inverse relation with birth weight over 

the third trimester (6.0-g decrease; [95% CI: -30.8, 18.8] per 10 ppb). Lin et al. (2004, 089503) 

reported a positive, though not statistically significant, exposure-response relationship for O3 during 

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1 the entire pregnancy in a Taiwanese study. In a study performed in Korea, Ha et al. (2001, 019390) 

2 reported no O3 effect during the first trimester of pregnancy, but they found that during the third 

3 trimester of pregnancy O3 was associated with LBW (RR 1.05 [95% CI: 1.02, 1.08] per 10 ppb). 

Table 7-2. Brief summary of epidemiologic studies of birth weight 

Study Location 

Sample Size Mean O3 (ppb) Exposure assessment Effect Estimate (95% CI) 

Salam et al. (2005, 087885) 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,

670076) 

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) 

Chen et al. (2002, 024945) 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,

088668) 

Los Angeles County, CA

(n=136,134) 1-h: 21.1-22.2 Varying distances frommonitor 

T1: NR 

T3: NR 

6 weeks before birth: NR 

Brauer et al. (2008,

156292) 

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,

088681) 

Nova Scotia, Canada 

(n=74,284) 24-h avg: 21 Nearest Monitor (within 25 km) 

T1: 0.97 (0.81, 1.18)

T2: 1.06 (0.87, 1.27)

T3: 1.01 (0.83-1.24) 

Mannes et al. (2005,

087895) 

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,

090703) 

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,

055613) 

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. (2004, 089503) 

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)^

T1: 1.02 (0.85, 1.22)^

T2: 0.93 (0.78, 1.12)^

T3: 1.05 (0.87, 1.26)^ 

Ha et al. (2001, 019390) Seoul, Korea 

(n=276,763) 

8-h avg: 22.4­

23.3* City-wide avg T1: 0.87 (0.81, 0.94)+T3: 1.05 (1.02, 1.08)+ 

*Median 

# Change in birthweight per 10 ppb change in O3 

^Odds ratios of LBW; Highest quartile of exposure compared to lowest quartile of exposure

+Relative risk of LBW per 10 ppb change in O3 

T1 = First Trimester, T2 = Second Trimester, T3 = Third Trimester 

NR: No quantitative results reported 

4 Table 7-2 provides a brief overview of the epidemiologic studies of birth weight. In summary, 

5 only the Children’s Health Study conducted in southern California (Salam et al., 2005, 087885) 

6 provides strong evidence for an effect of ambient O3 on birth weight. The study by Morello-Frosch 

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1 et al. (2010, 670076), also conducted in California, provides support for the results of the Children’s 

2 Health Study. Additional studies conducted in the U.S., Canada, Australia, Latin America, and Asia 

3 provide limited and inconsistent evidence to support the effect reported in the Children’s Health 

4 Study. The toxicological literature on the effect of O3 on birth weight is sparse. In some studies, the 

reporting of birth weight may be avoided because birth weight can be confounded by decreased litter 

6 size resulting from an increased rate of pup resorption (aborted pups) in O3 exposed dams. In one 

7 toxicological study by Haro and Paz (1993, 044194), no differences in litter size were observed and 

8 decreased birth weight in pups from dams who were exposed to 1ppm O3 during pregnancy was 

9 reported. 

7.4.4. Preterm Birth 

Preterm birth (PTB) is a syndrome (Romero et al., 2006, 625253) that is characterized by 

11 multiple etiologies. It is therefore unusual to be able to identify an exact cause for each PTB. In 

12 addition, PTB is not an adverse outcome in itself, but an important determinant of health status (i.e., 

13 neonatal morbidity and mortality). Although some overlap exists for common risk factors, different 

14 etiologic entities related to distinct risk factor profiles and leading to different neonatal and 

postneonatal complications are attributed to PTB and measures of fetal growth. Although both 

16 restricted fetal growth and PTB can result in LBW, prematurity does not have to result in LBW or 

17 growth restricted babies. 

18 A major issue in studying environmental exposures and preterm birth is selecting the relevant 

19 exposure period, since the biological mechanisms leading to preterm birth and the critical periods of 

vulnerability are poorly understood (Bobak, 2000, 011448). Exposures proximate to the birth may be 

21 most relevant if exposure causes an acute effect. However, exposure occurring in early gestation 

22 might affect placentation, with results observable later in pregnancy, or cumulative exposure during 

23 pregnancy may be the most important determinant. The studies reviewed have dealt with this issue in 

24 different ways. Many have considered several exposure metrics based on different periods of 

exposure. Often the time periods used are the first month (or first trimester) of pregnancy and the 

26 last month (or 6 weeks) prior to delivery. Using a time interval prior to delivery introduces an 

27 additional problem since cases and controls are not in the same stage of development when they are 

28 compared. For example, a preterm infant delivered at 36 weeks is a 32-week fetus 4 weeks prior to 

29 birth, while an infant born at term (40 weeks) is a 36-week fetus 4 weeks prior to birth. 

Recently, investigators have examined the association of PTB with both short- and long-term 

31 exposure periods. Time-series studies have been used to examine the association between air 

32 pollution concentrations during the days immediately preceding birth. An advantage of these time33 series studies is that this approach can remove the influence of covariates that vary across individuals 

34 over a short period of time. Retrospective cohort and case-control studies have been used to examine 

long-term exposure periods, often averaging air pollution concentrations over months or trimesters 

36 of pregnancy. 

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Reported studies fail to show consistency in pollutants and periods during pregnancy where an 

effect occurs. For example, while some studies find the strongest effects associated with exposures 

early in pregnancy, others report effects when the exposure is limited to the second or third trimester. 

However, the effect of air pollutant exposure during pregnancy on PTB has a biological basis. There 

is an expanding list of possible mechanisms that may explain the association between O3 exposure 

and PTB. These include: decreased in utero oxygen supply leading to a reduction of oxygen carrying 

capacity; changes in blood viscosity and disturbances of uterine blood flow; genetic mutations in 

first trimester leading to placental abnormalities; complex vascular alterations leading to placental 

abnormalities; disrupted implantation and placentation and suboptimal placental function; acute or 

sustained inflammatory response; disturbances of the pituitary-adrenocortico-placental system; and 

increased maternal susceptibility to infections. 

Many studies of PTB compare exposure in quartiles, using the lowest quartile as the reference 

(or control) group. No studies use a truly unexposed control group. If exposure in the lowest quartile 

confers risk, than it may be difficult to demonstrate additional risk associated with a higher quartile. 

Thus negative studies must be interpreted with caution. 

Preterm birth occurs both naturally (idiopathic preterm), and as a result of medical 

intervention (iatrogenic preterm). Ritz et al. (2000, 012068; 2007, 096146) excluded all births by 

Cesarean section to limit their studies to idiopathic preterm. No other studies attempted to 

distinguish the type of preterm birth, although air pollution exposure maybe associated with only one 

type. This is a source of potential effect misclassification. 

A number of air pollution-birth outcome studies have investigated the possible association 

between PTB and maternal exposure to O3. Most recently, Darrow et al. (2009, 195818) used vital 

record data to construct a retrospective cohort of 476,489 births occurring between 1994 and 2004 in 

5 central counties of metropolitan Atlanta. Using a time-series approach, the authors examined 

aggregated daily counts of preterm birth in relation to ambient levels of CO, NO2, SO2, O3, PM10, 

PM2.5 and speciated PM measurements. This study investigated 3 gestational windows of exposure: 

the first month of gestation, the final week of gestation, and the final 6 weeks of gestation. The 

authors did not observe associations of preterm birth with O3. 

A number of U.S. studies were conducted in southern California, and report somewhat 

inconsistent results. Ritz et al. (2000, 012068) evaluated the effect of air pollution (CO, NO2, O3, 

PM10) exposure during pregnancy on the occurrence of PTB in a cohort of 97,518 neonates born in 

southern California between 1989 and 1993. The authors averaged pollutant measures taken at the 

closest air-monitoring station over distinct periods, such as 1, 2, 4, 6, 8, 12, and 26 weeks before 

birth and the whole pregnancy period. Additionally, they calculated average exposures for the first 

and second months of pregnancy. The authors found no consistent effects for O3 over any of the 

pregnancy periods in single or multi-pollutant models. Wilhelm and Ritz (2005, 088668) extended 

previous analyses of PTB (Ritz and Yu, 1999, 086976; Ritz et al., 2000, 012068) in California to 

include 1994-2000. The authors examined varying residential distances from monitoring stations to 

see if the distance affected risk estimation, because effect attenuation may result from local pollutant 

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heterogeneity inadequately captured by ambient monitors. The authors analyzed the association 

between O3 exposure during varying periods of pregnancy and PTB, finding a positive association 

between O3 levels in both the first trimester of pregnancy (RR 1.23 [95% CI: 1.06, 1.42] per 10 ppb 

increase in 24-h avg O3) and the first month of pregnancy (results for first trimester exposure were 

similar, but slightly smaller, quantitative results not presented) in models containing all pollutants. 

No association was observed between O3 in the 6 weeks before birth and preterm delivery. Finally, 

Ritz et al. (2007, 096146) conducted a case-control survey nested within a birth cohort and assessed 

the extent to which residual confounding and exposure misclassification impacted air pollution effect 

estimates. The authors calculated mean exposure levels for three gestational periods: the entire 

pregnancy, the first trimester, and the last 6 weeks before delivery. Though positive associations 

were observed for CO and PM2.5, no consistent patterns of increase in the odds of preterm birth for 

O3 or NO2 were observed. 

One study conducted in Canada evaluated the impacts of air pollution (including CO, NO2, 

NO, O3, SO2, PM2.5, and PM10) on preterm births (1999-2002) using spatiotemporal residential 

exposure metrics by month of pregnancy in Vancouver, BC (Brauer et al., 2008, 156292). The 

authors did not observe consistent associations with any of the pregnancy average exposure metrics 

except for PM2.5 for PTB. The O3 associations were largely protective, and no quantitative results 

were presented for O3. Additionally, Lee et al. (2008, 195720) used time-series techniques to 

investigate the short-term associations of O3 and PTB in London, England. In addition to exposure 

on the day of birth, cumulative exposure up to 1 week before birth was investigated. The risk of 

preterm birth did not increase with exposure to the levels of ambient air pollution experienced by 

this population. 

Conversely, two studies conducted in Australia and one from China do provide evidence for an 

association between ambient O3 and PTB. Hansen et al. (2006, 089818) reported that exposure to O3 

during the first trimester was associated with an increased risk of PTB (OR 1.38, [95% CI: 

1.14, 1.69] per 10 ppb increase). Although the test for trend was significant due to the strong effect 

in the highest quartile, there was not an obvious exposure-response pattern across the quartiles of O3 

during the first trimester. The effect estimate was diminished and lost statistical significance when 

PM10 was included in the model (OR 1.23, [95% CI: 0.97, 1.59] per 10 ppb increase). Maternal 

exposure to O3 during the 90 days prior to birth showed a weak, positive association with PTB (OR 

1.09, [95% CI: 0.85, 1.39] per 10 ppb increase). Jalaludin et al. (2007, 156601) found that O3 levels 

in the month and three months preceding birth had a statistically significant association with PTB. 

Ozone levels in the first trimester of pregnancy were associated with increased risks for PTBs (OR 

1.15 [95% CI: 1.05, 1.24] per 10 ppb increase in 1-h max O3 concentration), and remained a 

significant predictor of preterm birth in co-pollutant models (ORs between 1.07 and 1.10). ORs 

increased for first month of pregnancy when restricted to within 5 km of a monitoring station (OR 

1.60, [95% CI: 1.27, 2.03]), but did not show a cumulative effect for first 3 months of pregnancy 

(OR 0.81, [95% CI: 0.67, 0.98]). Jiang et al. (2007, 093029) examined the acute effect of air 

pollution on preterm birth, including risk in relation to levels of pollutants for a single day exposure 

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1 window with lags from 0 to 6 days before birth. An increase of 10 ppb of the 8-week average of O3 

2 corresponded to 9.47 % (95% CI: 0.70, 18.7%) increase in PTBs. Increases in PTB were also 

3 observed for PM10, SO2, and NO2. The authors did not observe any significant acute effect of 

4 outdoor air pollution on PTB among the 1-day acute time windows examined in the week before 

5 birth. 

6 Little data is available from toxicological studies; one study reported a nearly statistically 

7 significant increase in pregnancy duration in mice when exposed to 0.8 or 1.2 ppm O3. This 

8 phenomenon was most likely due to the anorexigenic effect of relatively high O3 concentrations 

9 (Bignami et al., 1994, 076063). 

Table 7-3. Brief summary of epidemiologic studies of PTB 

Study 

Darrow et al. (2009,

195818) 

Location 

Sample Size 

Atlanta, GA 

(n=476,489) 

Mean O3 

(ppb) 

8-h max: 44.1 

Exposure 

assessment 

Population-weighted 

spatial averages

Nearest Monitor 

(within 4 miles) 

Effect Estimate# (95% CI) 

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,

012068) 

California, US 

(n=97,158) 8 h: 36.9 <2 mi of monitor First month: NR Last 6 weeks: NR 

Wilhelm and Ritz 

(2005, 088668) 

Ritz et al. (2007,

096146) 

Brauer et al. (2008,

156292) 

Lee et al. (2008, 

195720) 

Hansen et al. (2006,

089818) 

Los Angeles, CA

(n=106,483) 

Los Angeles, CA

(n=58,316) 

Vancouver, BC, 

Canada 

(n=70,249) 

London, UK 

Brisbane, 

Australia 

(n=28,200) 

1 h: 21.1-22.2 Varying distances to

monitor 

24-h avg:

22.5 

Nearest monitor to 

ZIP code 

24-h avg: 14 

Nearest Monitor 

(within 10 km)

Inverse Distance 

Weighting (IDW) 

24-h avg: NR 1 monitor 

8-h max: 26.7 City-wide avg 

First month: 1.23 (1.06, 1.42) 

T1: NR 

T2: 1.38 (1.14, 1.66) 

Last 6 weeks: NR 

Entire pregnancy: NR 

T1: 0.93 (0.82, 1.06) 

Last 6 weeks: NR 

Entire pregnancy: NR

First 30 days of pregnancy: NR

Last 30 days of pregnancy: NR

T1: NR 

T3: NR 

Lag 0: 1.00 (1.00, 1.01) 

T1: 1.39 (1.15, 1.70) 

T3: 1.09 (0.88, 1.39) 

Jalaludin et al. (2007,

156601) 

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)* 

T1: 0.807 (0.668, 0.976)* 

T3: 1.011 (0.910, 1.124)* 

Last month: 0.984 (0.906, 1.069)* 

Shanghai, China Jiang et al. (2007, (n=3,346 preterm 8-h avg: 32.7 City-wide avg 093029) births) 

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) 

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#Relative risk of PTB per 10 ppb change in O3.
 

*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 7-3 provides a brief overview of the epidemiologic studies of PTB. In summary, the 

2	 evidence is consistent when examining shorter-term, late-pregnancy exposure to O3 and reports no 

3	 association with PTB. However when long-term exposure to O3 early in pregnancy is examined the 

4	 results are inconsistent. Studies conducted in the U.S., Canada, and England find no association with 

O3 and PTB, while studies conducted in Australia and China report an O3 effect on PTB. 

6 

7.4.5. Fetal Growth 

7 Low birth weight has often been used as an outcome measure because it is easily available and 

8 accurately recorded on birth certificates. However, LBW may result from either short gestation, or 

9 inadequate growth in utero. Most of the studies investigating air pollution exposure and LBW 

limited their analyses to term infants to focus on inadequate growth. A number of studies were 

11 identified that specifically addressed growth restriction in utero by identifying infants who failed to 

12 meet specific growth standards. Usually these infants had birth weight less than the 10th percentile 

13 for gestational age, using an external standard. Many of these studies have been previously 

14 discussed, since they also examined other reproductive outcomes (i.e., LBW or PTB). 

A limitation of environmental studies that use birth weight as a proxy measure of fetal growth 

16 is that patterns of fetal growth during pregnancy cannot be assessed. This is particularly important 

17 when investigating pollutant exposures during early pregnancy as birth weight is recorded 

18 many months after the exposure period. The insult of air pollution may have a transient effect on 

19 fetal growth, where growth is hindered at one point in time but catches up at a later point. For 

example, maternal smoking during pregnancy can alter the growth rate of individual body segments 

21 of the fetus at variable developmental stages, as the fetus experiences selective growth restriction 

22 and augmentation (Lampl and Jeanty, 2003, 625252). 

23 Fetal growth is influenced by maternal, placental, and fetal factors. The biological mechanisms 

24 by which air pollutants may influence the developing fetus remain largely unknown. Several 

mechanisms have been proposed, including maternal susceptibility to infection, oxidative stress, 

26 hematological factors such as blood viscosity, and the direct effect of specific pollutants on fetal 

27 development or on DNA and its transcription. Air pollution may affect maternal respiratory function 

28 or general health, which may in turn impair uteroplacental and umbilical blood flow, transplacental 

29 glucose, and total insulin, all of which are important determinants of fetal growth. Additionally, 

certain changes resulting in fetal growth retardation may occur in early pregnancy (around the time 

31 of implantation) caused by an abnormal reaction between the trophoblast and uterine tissues. A 

32 defective trophoblast invasion, resulting in suboptimal placentation and maternal hemodynamic 

33 maladaptation can alter growth and development of the fetus. Inhalation of air pollution can cause 

34 inflammatory responses and oxidative stress, and both of these reactions can interfere with normal 

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intrauterine growth via vascular dysfunction in the placenta and damaged DNA. Also, proinflammatory cytokines can limit trophoblast invasion during the early stages of pregnancy, 

restricting fetal growth. Poor placental vascularity is caused partly by dysregulation of gene 

expression in key angiogenic factors in early pregnancy, and if ambient air pollution is associated 

with poor placental function it may partly be caused by perturbed DNA transcription early in 

pregnancy. 

The terms small-for-gestational-age (SGA), which is defined as a birth weight <10th percentile 

for gestational age (and often sex and/or race), and intrauterine growth retardation (IUGR) are often 

used interchangeably. However, this definition of SGA does have limitations. For example, using it 

for IUGR may overestimate the percentage of “growth-restricted” neonates as it is unlikely that 10% 

of neonates have growth restriction (Wollmann, 1998, 193812). On the other hand, when the 10th 

percentile is based on the distribution of live births at a population level, the percentage of SGA 

among PTB is most likely underestimated (Hutcheon and Platt, 2008, 193795). Nevertheless, SGA 

represents a statistical description of a small neonate, whereas the term IUGR is reserved for those 

with clinical evidence of abnormal growth. Thus all IUGR neonates will be SGA, but not all SGA 

neonates with be IUGR (Wollmann, 1998, 193812). In the following section the terms SGA and 

IUGR are referred to as each cited study used the terms. 

Over the past decade a number of studies examined various metrics of fetal growth restriction. 

Salam et al. (2005, 087885) assessed the effect of increasing O3 concentrations on IUGR in a 

population of infants born in California from 1975-1987 as part of the Children’s Health Study. The 

authors reported that maternal O3 exposures averaged over the entire pregnancy and during the third 

trimester were associated with increased risk of IUGR. A 10-ppb difference in 24-h maternal O3 

exposure during the third trimester increased the risk of IUGR by 11% (95% CI: 0, 20%). Brauer et 

al. (2008, 156292) evaluated the impacts of air pollution (CO, NO2, NO, O3, SO2, PM2.5, PM10) on 

SGA (1999-2002) using spatiotemporal residential exposure metrics by month of pregnancy in 

Vancouver, BC. The O3 associations were largely protective (OR= 0.87, [95% CI: 0.81, 0.93] for a 

10 ppb increase in inverse distance weighted SGA), and no additional quantitative results were 

presented for O3. Liu et al. (2007, 090429) examined the association between IUGR among singleton 

term live births and SO2, NO2, CO, O3, and PM2.5 in 3 Canadian cities for the period 1985-2000. No 

increase in the risk of IUGR in relation to exposure to O3 averaged over each month and trimester of 

pregnancy was noted. 

Three studies conducted in Australia provide evidence for an association between ambient O3 

and fetal growth restriction. Hansen et al. (2007, 090703) examined SGA among singleton, full-term 

births in Brisbane, Australia in relation to ambient air pollution (bsp, PM10, NO2, O3) during 

pregnancy. They also examined head circumference and crown-heel length in a subsample of term 

neonates. Trimester specific exposures to all pollutants were not statistically significantly associated 

with a reduction in head circumference or an increased risk of SGA. When monthly specific 

exposures were examined, the authors observed an increased risk of SGA associated with exposure 

to O3 during month 4 (OR 1.11 [95% CI: 1.00, 1.24] per 10 ppb increase). In a subsequent study, 

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Hansen et al. (2008, 190273) examined the possible associations between fetal ultrasonic 

measurements and ambient air pollution (PM10, O3, NO2, SO2) during early pregnancy. This study 

had two strengths: (1) fetal growth was assessed during pregnancy as opposed to at birth; and (2) 

there was little delay between exposures and fetal growth measurements, which reduces potential 

confounding and uses exposures that are concurrent with the observed growth pattern of the fetus. 

Fetal ultrasound biometric measurements were recorded for biparietal diameter (BPD), femur length, 

abdominal circumference, and head circumference. To further improve exposure assessment, the 

authors restricted the samples to include only scans from women for whom the centroid of their 

postcode was within 14 km of an air pollution monitoring site. Ozone during days 31-60 was 

associated with decreases in all of the fetal growth measurements, and a 1.78 mm reduction in 

abdomen circumference per 10 ppb increase in O3 concentration, though this effect did not persist in 

co-pollutant models. The change in ultrasound measurements associated with O3 during days 31-60 

of gestation indicated that increasing O3 concentration decreased the magnitude of ultrasound 

measurements for women living within 2 km of the monitoring site. The relationship decreased 

toward the null as the distance from the monitoring sites increased. When assessing effect 

modification due to SES, there was some evidence of effect modification for most of the 

associations, with the effects of air pollution stronger in the highest SES quartile. In the third study, 

Mannes et al. (2005, 087895) estimated the effects of pollutant (PM10, PM2.5, NO2, CO and O3) 

exposure in the first, second and third trimesters of pregnancy and risk of SGA in Sydney, Australia. 

Citywide average air pollutant concentrations in the last month, third trimester, and first trimester of 

pregnancy had no effect on SGA. Concentrations of O3 in the second trimester of pregnancy had 

small but adverse effects on SGA (OR 1.10 [95% CI: 1.00, 1.14] per 10 ppb increment). This effect 

disappeared when the analysis was limited to births with a maternal address within 5 km of a 

monitoring station (OR 1.00 [95% CI: 0.60, 1.79] per 10 ppb increment). 

Very little information from toxicological studies is available to address effects on fetal 

growth. However, there is evidence to suggest that prenatal exposure to O3 can affect postnatal 

growth. A few studies reported that mice or rats exposed developmentally (gestationally ± 

lactationally) to O3 had deficits in body weight gain in the postpartum period (Bignami et al., 1994, 

076063; Haro and Paz, 1993, 044194; Kavlock et al., 1980, 094043). 

Table 7-4 provides a brief overview of the epidemiologic studies of fetal growth restriction. In 

summary, the evidence is inconsistent when examining exposure to O3 and fetal growth restriction. 

Similar to PTB, studies conducted in Australia have reported an effect of O3 on fetal growth, whereas 

studies conducted in other areas have not found such an effect. This may be due to the restriction of 

births to those within 2-14 km of a monitoring station, as was done in the Australian studies. 

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Table 7-4. Brief summary of epidemiologic studies of fetal growth 

Location	 Exposure Study	 Mean O3 (ppb) Effect Estimate (95% CI) (Sample Size)	 assessment 

Entire pregnancy: 1.16 (1.00, 1.32) 

Salam et al. (2005, California, U.S. 24-h avg: 27.3 T1:1.00 (0.94, 1.11) ZIP code level 087885) (n=3901) 8 h: 50.6 T2: 1.06 (1.00, 1.12) 

T3: 1.11 (1.00, 1.17) 

Entire pregnancy: NR Nearest Monitor (within First 30 days of pregnancy: NR Brauer et al. (2008, Vancouver, BC, Canada	 10 km) 24-h avg: 14	 Last 30 days of pregnancy: NR 156292) (n=70,249)	 Inverse Distance T1: NR Weighting (IDW) T3: NR 

Entire pregnancy: NR (results

Calgary, Edmonton, and presented in figure) Liu et al. (2007, 24-h avg: 16.5 Census Subdivision Montreal, Canada	 T1: NR (results presented in figure) 090429)	 1-h max: 31.2 avg (n= 16,430)	 T2: NR (results presented in figure)

T3: NR (results presented in figure) 

T1: 1.01 (0.89, 1.15) Hansen et al. (2007, Brisbane, Australia 8-h max: 26.7 City-wide avg T2: 1.00 (0.86, 1.17) 090703) (n=26,617) 

T3: 0.83 (0.71, 0.97) 

M1: -0.32 (-1.56, 0.91)*

Hansen et al. (2008, Brisbane, Australia M2: -0.58 (-1.97, 0.80)* 8-h avg: 24.8 Within 2 km of monitor 190273) (n=15,623)	 M3: 0.26 (-1.07, 1.59)*

M4: 0.11 (-0.98, 1.21)* 

T1: 0.90 (0.48, 1.34) 

T2: 1.00 (0.60, 1.79) Mannes et al. (2005, Sydney, Australia City-wide avg and 1-h max: 31.6 087895) (n=138,056) <5 km from monitor	 T3: 1.10 (0.66, 1.97) 

Last 30 days of pregnancy: 1.10 

(0.74, 1.79) 

#Relative risk of fetal growth restriction per 10 ppb change in O3. 

*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. 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 effect of 

3 temporal variations in ambient air pollution on birth defects. Heart defects and oral clefts have been 

4 the focus of the majority of these recent studies, given the higher prevalence than other birth defects 

5 and associated mortality. 

6 Mechanistically, air pollutants could be involved in the etiology of birth defects via 

7 hemodynamic events, anoxic events, oxidative stress, and toxicity to certain cell populations during 

8 development. One potential etiologic pathway may include the neural crest cell population. Normal 

9 migration and differentiation of neural crest cells are important for heart development and are 

10 particularly sensitive to toxic insults. They respond by undergoing apoptosis, in part because they 

11 lack antioxidative stress proteins. Ozone is a very reactive molecule and a strong oxidizing agent that 

12 can generate superoxides, hydrogen peroxide, and hydroxyl radicals, contributing to oxidative stress 

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Several studies have been conducted examining the relationship between O3 exposure during 

pregnancy and birth defects. The earliest of these studies was conducted in southern California (Ritz 

et al., 2002, 023227). This study evaluated the effect of air pollution on the occurrence of cardiac 

and orofacial birth defects in neonates and fetuses delivered in southern California in 1987-1993. 

Maternal exposure estimates were based on data from the fixed site closest to the mother’s ZIP code 

area. When using a case-control design where cases were matched to 10 randomly selected controls, 

results showed increased risks for aortic artery and valve defects (OR 1.56 [95% CI: 1.16, 2.09] per 

10 ppb O3), pulmonary artery and valve anomalies (OR 1.34 [95% CI: 0.96, 1.87] per 10 ppb O3), 

and conotruncal defects (OR 1.36 [95% CI: 0.91, 2.03] per 10 ppb O3) in a dose-response manner 

with second-month O3 exposure. These associations were not observed for orofacial defects. The 

average effect sizes and patterns of second-month O3 exposure were similar for these defects and 

varied only slightly from single- to multi-pollutant models, or when the models were adjusted for 

other potential confounding factors. Additionally, the authors reported an OR of 1.13 (95% CI: 0.90, 

1.40) per 10 ppb during the second trimester for cleft lip with or without cleft palate. 

A study conducted in Texas (Gilboa et al., 2005, 087892) looked at a similar period of 

exposure but reported no association with most of the birth defects studied (O3 concentration was 

studied using quartiles with the lowest representing <18 ppb and the highest representing ≥ 31 ppb). 

The authors found slightly elevated odds ratios for pulmonary artery and valve defects. They also 

detected an inverse association between O3 exposure and isolated ventricular septal defects. For cleft 

lip with or without cleft palate, the authors reported and OR of 1.09 (95% CI: 0.70, 1.69) for the 

fourth quartile contrasted with the first quartile of exposure during 3-8 weeks of pregnancy. Overall, 

this study did not provide strong evidence that air pollution increases the risk of cardiac defects or 

oral clefts. 

A recent study conducted in Atlanta, GA examined O3 exposure during the third through 

seventh week of pregnancy and reported no association with risk of cardiovascular malformations 

(mean long-term average of 8-h O3 concentrations excluding November through February ranged by 

5-year groups from 39.8 to 43.3 ppb) (Strickland et al., 2009, 190324). 

Hwang and Jaakola (2008, 193794) conducted a population-based case-control study to 

investigate exposure to ambient air pollution and the risk of cleft lip with or without cleft palate in 

Taiwan. The risk of cleft lip with or without cleft palate was increased in relation to O3 levels in the 

first gestational month (OR 1.17 [95% CI: 1.01, 1.36] per 10 ppb) and second gestational month 

(OR 1.22 [95% CI: 1.03, 1.46] per 10 ppb), but was not related to any of the other pollutants. In 

three-pollutant models, the effect estimates for O3 exposure were stable for the four different 

combinations of pollutants and were all statistically significant. 

Marshall et al. (2010, 597374) compared estimated exposure to ambient pollutants during 

early pregnancy among mothers of children with oral cleft defects to that among mothers of controls. 

The authors observed no consistent elevated associations between any of the air pollutants examined 

and cleft malformations, though there was a weak association between cases of cleft palate only and 

increasing O3 concentrations. This association increased when cases and controls were limited to 

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1 those with residences within 10 km of the closest O3 monitor (OR 2.2 [95% CI: 1.0, 4.9], comparing 

2 highest quartile [>33 ppb] to lowest quartile [<15 ppb]). 

3 A limited number of toxicological studies have examined birth defects in animals exposed 

4 gestationally to O3. Kavlock et al. (1979, 039228) exposed pregnant rats to O3 for precise periods 

5 during organogenesis. No significant teratogenic effects were found in rats exposed 8 hr/day to 

6 concentrations of O3 varying from 0.44 to 1.97 ppm during early (days 6-9), mid (days 9-12), or late 

7 (days 17 to 20) gestation, or the entire period of organogenesis (days 6-15). Earlier research found 

8 eyelid malformation following gestational and postnatal exposure to 0.2 ppm O3 (Veninga, 1967, 

9 040746). 

10 Table 7-5 provides a brief overview of the epidemiologic studies of birth defects. Results from 

11 these studies are not entirely consistent. This inconsistency could be due to the absence of true 

12 associations between O3 and risks of cardiovascular malformations and oral cleft defects; it could 

13 also be due to differences in populations, pollution levels, outcome definitions, or analytical 

14 approaches. The lack of consistency of associations between O3 and cardiovascular malformations or 

15 oral cleft defects might be due to issues relating to statistical power or measurement error. 

Table 7-5. 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,

023227) 

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,

087892) 

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, 193794) 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, 190324) Cardiac Defects Atlanta, GA (n=3,338 cases) 8-h max: 39.8-43.3 Weighted City-wide avg 

Weeks 3-7 of 

gestation 

Marshall et al. (2010,

597374) 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 

7.4.7. De velop me n tal Res piratory Effects 

16 The issue of prenatal exposure has assumed increasing importance, since ambient air pollution 

17 exposures of pregnant women have been shown to lead to adverse pregnancy outcomes, as well as to 

18 respiratory morbidity and mortality in the first year of life. Growth and development of the 

19 respiratory system take place mainly during the prenatal and early postnatal periods. This early 

20 developmental phase is thought to be very important in determining long-term lung growth. Studies 

21 have recently examined this emerging issue, and several were included in Sections 7.2.1 and 7.2.3, 

22 and are included here because they included both prenatal and post-natal exposure periods. 

23 Mortimer et al. (2008, 122163; 2008, 187280) examined the association of prenatal and 

24 lifetime exposures to air pollutants with pulmonary function and allergen sensitization in a subset of 

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

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

separately across several important developmental time-periods, including the entire pregnancy, each 

trimester, the first 3 years of life, the first 6 years of life, and the entire lifetime. The 8-h avg O3 

concentrations were approximately 50 ppb for each of the exposure metrics (estimated from figure). 

In the first analysis (Mortimer et al., 2008, 122163), negative effects on pulmonary function were 

found for exposure to PM10, NO2, and CO during key neonatal and early life developmental periods. 

The authors did not find a negative effect of exposure to O3 among this cohort. In the second 

analysis (Mortimer et al., 2008, 187280), sensitization to at least one allergen was associated, in 

general, with higher levels of CO and PM10 during the entire pregnancy and second trimester and 

higher PM10 during the first 2 years of life. Lower exposure to O3 during the entire pregnancy or 

second trimester was associated with an increased risk of allergen sensitization. Although the 

pollutant metrics across time periods are correlated, the strongest associations with the outcomes 

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

and postnatal exposures, the models from this group of studies suggest that each time period of 

exposure may contribute independently to different dimensions of school-aged children’s pulmonary 

function. For 4 of the 8 pulmonary-function measures (FVC, FEV1, PEF, FEF25-75), prenatal 

exposures were more influential on pulmonary function than early-lifetime metrics, while, in 

contrast, the ratio of measures (FEV1/FVC and FEF25-75/FVC) were most influenced by postnatal 

exposures. When lifetime metrics were considered alone, or in combination with the prenatal 

metrics, the lifetime measures were not associated with any of the outcomes, suggesting the timing 

of the exposure may be more important than the overall dose and prenatal exposures are not just 

markers for lifetime or current exposures. 

Clark et al. (2010, 594440) investigated the effect of exposure to ambient air pollution in utero 

and during the first year of life on risk of subsequent asthma diagnosis (incident asthma diagnosis up 

to age 3-4) in a population-based nested case-control study. Air pollution exposure for each subject 

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

regression modeling, and proximity to stationary pollution sources. An average exposure was 

calculated for the duration of pregnancy (~15 ppb; transformed from μg/m3

) and the first year of life 

(~14 ppb; transformed from μg/m3). In contrast to the Mortimer et al. studies (2008, 122163; 2008, 

187280), the effect estimates for first-year exposure were generally larger than for in utero 

exposures. However, similar to the Mortimer et al. studies, the observed associations with O3 were 

largely protective. Because of the relatively high correlation between in utero and first-year 

exposures for many pollutants, it was difficult to discern the relative importance of the individual 

exposure periods. 

Latzin et al. (2009, 195721) examined whether prenatal exposure to air pollution was 

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

inhomogeneity and eNO were measured in 241 unsedated, sleeping neonates (age= 5 weeks). The 

median of the 24-h avg O3 concentrations averaged across the post-natal period was ~44 ppb 

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(transformed from μg/m3). Consistent with the previous studies, no association was found for 

prenatal exposure to O3 and lung function. 

The new toxicological literature since the 2006 O3 AQCD (U.S. EPA, 2006, 088089), covering 

respiratory changes related to developmental O3 exposure, reports ultrastructural changes in 

bronchiole development, alterations in placental and pup cytokines, and increased pup airway hyperreactivity. These studies are detailed below. Older studies are discussed where new information is not 

available. 

Fetal rat lung bronchiole development is triphasic, comprised of the glandular phase 

(measured at GD18), the canalicular phase (GD20), and the saccular phase (GD21). The 

ultrastructural lung development in fetuses of pregnant rats exposed to 1-ppm O3 (12 h/day, out to 

either GD18, GD20 or GD21) was examined by electron microscopy during these three phases. In 

the glandular phase, bronchiolar columnar epithelial cells in fetuses of dams exposed to O3 had 

cytoplasmic damage and swollen mitochondria. Bronchial epithelium at the canalicular phase in O3 

exposed pups had delayed maturation in differentiation, i.e., glycogen abundance in secretory cells 

had not diminished as it should with this phase of development. Congruent with this finding, delayed 

maturation of tracheal epithelium following early neonatal O3 exposure (1 ppm, 4-5 h/day for 

first week of life) in lambs has been previously reported (Mariassy et al., 1989, 042246; Mariassy et 

al., 1990, 042311). Also at the canalicular phase, atypical cells were seen in the bronchiolar lumen of 

O3 exposed rat fetuses. Finally, in the saccular phase, mitochondrial degradation was present in the 

non-ciliated bronchiolar cells of rats exposed in utero to O3. In conclusion, O3 exposure of pregnant 

rats produced ultra-structural damage to near-term fetal bronchiolar epithelium (López et al., 2008, 

197786). 

Exposure of laboratory animals to multiple airborne pollutants can differentially affect pup 

physiology. One study showed that exposure of C57BL/6 mouse dams to 0.48 mg PM intratracheally 

twice weekly for 3 weeks during pregnancy augmented O3-induced airway hyper-reactivity in 

juvenile offspring. Maternal PM exposure also significantly increased placental cytokines above 

vehicle-instilled controls. Pup postnatal O3 exposure (1 ppm 3 h/day, every other day, thrice weekly 

for 4 weeks) induced significantly increased cytokine levels (IL-1β, TNF-α, KC, and IL-6) in whole 

lung versus postnatal air exposed groups; this was further exacerbated with gestational PM exposure 

(Auten et al., 2009, 200760). 

A series of experiments using infant rhesus monkeys repeatedly exposed to 0.5 ppm O3 

starting at one-month of age have examined the effect of O3 alone or in combination with an inhaled 

allergen on morphology and lung function (Plopper et al., 2007, 596412). Exposure to O3 alone or 

allergen alone produced small but not statistically significant changes in baseline airway resistance 

and airway responsiveness, but the combined exposure to both O3 + antigen produced statistically 

significant and greater than additive changes in both functional measurements. Additionally, cellular 

changes and significant structural changes in the respiratory tract have been observed in infant 

rhesus monkeys exposed to O3 (Fanucchi et al., 2006, 096491). A more detailed description of these 

studies can be found in Section 7.2.3 (Pulmonary Structure and Function). 

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7.4.8. De velop me n tal Central Nervous Sys tem Effects 

7.4.8.1. Laterality 

1 Two reports of laterality changes in mice developmentally exposed to O3 have been reported in 

2 the literature. Mice developmentally exposed to 0.6 ppm O3 (6 days before breeding to weaning at 

3 PND21) showed a turning preference (left turns) distinct from air exposed controls (clockwise turns) 

4 (Dell'Omo et al., 1995, 080787); in previous studies this behavior in mice has been found to 

correlate with specific structural asymmetries of the hippocampal mossy fiber projections (Schöpke 

6 et al., 1991, 684254). The 2006 AQCD evidence for the effect of O3 on laterality or handedness 

7 demonstrated that rats exposed to O3 during fetal and neonatal life showed limited, gender-specific 

8 changes in handedness after exposure to the intermediate dose of O3 (only seen in female mice 

9 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 with 

exposure from 6 days before breeding to PND26) (Petruzzi et al., 1999, 012066). 

7.4.8.2. Bra in Morp hology a nd Ne uroc he mic a l Cha n g e s 

11 The nucleus tractus solitarius (NTS), a medullary area of respiratory control, of adult animals 

12 exposed prenatally to 0.5 ppm O3 (12h/day, ED5-ED20) had significantly less tyrosine hydroxylase 

13 staining versus control (Boussouar et al., 2009, 596368). Tyrosine hydroxylase is the rate-limiting 

14 enzyme for dopamine synthesis and serves as a precursor for catecholamine synthesis; thus, 

decreased staining is used as a marker of dopaminergic or catecholaminergic cell or activity loss in 

16 these regions and thus functions in neuronal plasticity. After physical restraint stress, control animals 

17 respond at the histological level with Fos activation, a marker of neuronal activity, and tyrosine 

18 hydroxylase activation in the NTS, a response which is absent or attenuated in adult animals exposed 

19 prenatally to 0.5 ppm O3 (Boussouar et al., 2009, 596368) when compared to control air exposed 

animals who also were restrained. The O3-exposed offspring in this study were cross-fostered to 

21 control air exposed dams to avoid O3-dependent dam related neonatal effects on offspring outcomes 

22 (i.e., dam behavioral or lactational contributions to pup outcomes) (Boussouar et al., 2009, 596368). 

23 Developmental exposure to 0.3 or 0.6 ppm O3 prior to mating pair formation through GD17 

24 induced significant increased levels of BDNF in the striatum of adult (PND140) O3 exposed 

offspring as compared to control air exposed animals; these O3-exposed animals also had 

26 significantly decreased level of NGF in the hippocampus versus control (Santucci et al., 2006, 

27 596414). 

28 Changes in the pup cerebellum with prenatal 1 ppm O3 exposure include altered morphology 

29 (Rivas-Manzano and Paz, 1999, 012072; Romero-Velazquez et al., 2002, 035575), decreased total 

area (Romero-Velazquez et al., 2002, 035575), decreased number of Purkinje cells (Romero31 Velazquez et al., 2002, 035575), and altered monoamine neurotransmitter content with the 

32 catecholamine system affected and the indoleamine system unaffected by O3 (Gonzalez-Pina et al., 

33 2008, 475317). 

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7.4.8.3. Ne urobe ha viora l Ou tc o me s 

1 

2 O3 administration to dams during pregnancy with or without early neonatal exposure has been 

3 shown to contribute to multiple neurobehavioral outcomes in offspring that are described in further 

4 detail below. 

O3 administration (0.4, 0.8 or 1.2 ppm O3) during the majority of pregnancy (PD7-17) of CD-1 

6 mice did not affect pup behavioral outcomes including early behavioral ultrasonic vocalizations and 

7 more permanent later measurements (PND60 or 61) including pup activity, habituation and 

8 exploration and d-amphetamine-induced hyperactivity (Bignami et al., 1994, 076063); these pups 

9 were all cross-fostered or reared on non- O3 exposed dams. 

Testing for aggressive behavior in mice continuously exposed to O3 (0.3 or 0.6 ppm from 

11 30 days prior to mating to GD17) revealed that mice had significantly increased defensive/ 

12 submissive behavior (increased freezing posturing on the first day only of a multiple-day exam) 

13 versus air exposed controls (Santucci et al., 2006, 596414). Similar to this and as reported in 

14 previous AQCDs, continuous exposure of adult animals to O3 induced significant increases in fear 

behavior and decreased aggression as measured by significantly decreased freezing behavior 

16 (Petruzzi et al., 1995, 077448). 

17 Developmentally exposed animals also had significantly decreased amount of time spent nose 

18 sniffing other mice (Santucci et al., 2006, 596414); this social behavior deficit, decreased sniffing 

19 time, was not found in an earlier study with similar exposures (Petruzzi et al., 1995, 077448), but 

sniffing of specific body areas was measured in Santucci et al. (2006, 596414) and total number of 

21 sniffs of the entire body was measured in Petruzzi et al. (1995, 077448). The two toxicology studies 

22 exploring social behavior (sniffing) employ different study designs and find opposite effects in 

23	 animals exposed to O3. 

7.4.8.4. Sle e p Abe rra tions a fte r De ve lop me n ta l Ozone Expos ure 

24 The effect of gestational O3 exposure (1 ppm O3, 12h/day, during dark period) on sleep 

patterns in rat offspring was followed using 24 h polysomnographic recordings at 30, 60 and 90 days 

26 of age (Haro and Paz, 1993, 044194). Ozone-exposed pups manifested with inverted sleep-wake 

27 patterns or circadian rhythm phase-shift. Rat vigilance was characterized in wakefulness, slow wave 

28	 sleep (SWS), and paradoxical sleep (PS) using previously characterized criteria. The O3 exposed 

29	 offspring spent longer time in the wakefulness state during the light period, more time in SWS 

during the period of darkness, and showed significant decrements in PS. Chronic O3 inhalation 

31 significantly decreased the duration of PS during both the light and dark periods (Haro and Paz, 

32 1993, 044194). These effects were consistent at all time periods measured (30, 60 and 90 days of 

33 age). These sleep effects reported after developmental exposures expand upon the existing literature 

34	 on sleep aberrations in adult animals exposed to O3 [rodents: (Arito et al., 1992, 042759; Paz and 

Huitron-Resendiz, 1996, 082684); and cats: (Paz and Bazan-Perkins, 1992, 036436)]. A role for 

36	 inhibition of cyclooxygenase-2 and the interleukins and prostaglandins in the O3-dependent sleep 

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1 changes potentially exists with evidence from a publication on indomethacin pretreatment
 

2 attenuating O3-induced sleep aberrations in adult male animals (Rubio and Paz, 2003, 053541). 

7.4.9. Early Life Mortality 

3 Infants may be particularly susceptible to the adverse effects of air pollution. The lung is not 

4 well developed at birth, with 80% of alveoli being formed postnatally. An important question 

regarding the association between PM and infant mortality is the critical window of exposure during 

6 development for which infants are susceptible. Several age intervals have been explored: neonatal 

7 (<1 month); postneonatal (1 month to 1 year); and an overall interval for infants that includes both 

8 the neonatal and postneonatal periods (<1 year). Within these various age categories, multiple causes 

9 of deaths have been investigated, particularly total deaths and respiratory-related deaths. The studies 

reflect a variety of study designs, exposure periods, regions, and adjustment for confounders. Within 

11 the first year of life, infants develop rapidly; therefore their susceptibility may change within weeks 

12 or months. During the neonatal and post-neonatal periods, the developing lung is highly susceptible 

13 to environmental toxicants. As discussed below, a handful of studies have examined the effect of 

14 ambient air pollution on neonatal and postneonatal mortality, with the former the least studied. These 

studies varied somewhat with regard to the outcomes and exposure periods examined and study 

16 designs employed. 

17 The results of these infant mortality studies are presented here, and in Table 7-6, with the other 

18 reproductive and developmental outcomes because it is likely that in utero exposures contribute to 

19 this outcome. Both long-term and short-term exposure studies of infant mortality are included in this 

section. A major issue in studying environmental exposures and infant mortality is selecting the 

21 relevant exposure period, since the biological mechanisms leading to death and the critical periods of 

22 vulnerability are poorly understood. Exposures proximate to the death may be most relevant if 

23 exposure causes an acute effect. However, exposure occurring in early life might affect critical 

24 growth and development, with results observable later in the first year of life, or cumulative 

exposure during the first year of life may be the most important determinant. The studies reviewed 

26 below have dealt with this issue in different ways. Many have considered several exposure metrics 

27 based on different periods of exposure. 

7.4.9.1. Stillbirth 

28 Pereira et al. (1998, 007264) investigated the association among daily counts of intrauterine 

29 mortality (over 28 weeks of gestation) and air pollutant concentrations in Sao Paulo, Brazil from 

1991 through 1992. The association was strong for NO2, but lesser for SO2 and CO. These 

31 associations exhibited a short lag time, less than 5 days. No significant association was detected 

32 between O3 and intrauterine mortality. 

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7.4.9.2. Infa nt Mo rtality, Les s than 1 Year 

1 Ritz et al. (2006, 089819) linked birth and death certificates for infants who died between 

2 1989 and 2000 to evaluate the influence of outdoor air pollution on infant death in the South Coast 

3 Air Basin of California. The authors examined exposure periods 2 weeks, 1 month, 2 months, and 

4 6 months before each case subject's death and reported no association between ambient levels of O3 

and infant mortality. Similarly, Diaz et al. (2004, 089894) analyzed the effects of extreme 

6 temperatures and air pollutants on daily mortality in children less than 1 year of age in Madrid, 

7 Spain, from 1986 to 1997 and observed no statistically significant association between mortality and 

8 O3 concentrations. Hajat et al. (2007, 093276) analyzed time-series data of daily infant mortality 

9 counts in 10 major cities in the UK to quantify any associations with short-term changes in air 

pollution. When the results from the 10 cities were combined there was no relationship between O3 

11 and infant mortality, even after restricting the analysis to just the summer months. 

12 Conversely, a time-series study of infant mortality conducted in the southwestern part of 

13 Mexico City in the years 1993-1995 found that infant mortality was associated with the levels of 

14 NO2 and O3 3-5 days before death, but not as consistently as with PM. A 10-ppb increase in 24-h avg 

O3 was associated with a 2.78% increase (95% CI: 0.29, 5.26%) in infant mortality (lag 3) (Loomis 

16 et al., 1999, 087288). This increase was attenuated, though still positive when evaluated in a two17 pollutant model with PM2.5. One-hour max concentrations of O3 exceeded prevailing Mexican and 

18 international standards nearly every day. 

7.4.9.3. Ne o na ta l Morta lity, Les s than 1 Month 

19	 Three studies have evaluated ambient O3 concentrations and neonatal mortality and observed 

no association. Ritz et al. (2006, 089819) linked birth and death certificates for infants who died 

21 between 1989 and 2000 to evaluate the influence of outdoor air pollution on infant death in the South 

22 Coast Air Basin of California. The authors examined exposure periods 2 weeks, 1 month, 2 months, 

23 and 6 months before each case subject's death and reported no association between ambient levels of 

24 O3 and neonatal mortality. Hajat et al. (2007, 093276) analyzed time-series data of daily infant 

mortality counts in 10 major cities in the UK to quantify any associations with short-term changes in 

26 air pollution. When the results from the 10 cities were combined there was no relationship between 

27 O3 and neonatal mortality, even after restricting the analysis to just the summer months. Lin et al. 

28 (2004, 095787) assessed the impact of daily changes in air pollutants on the number of daily 

29 neonatal deaths in Sao Paulo, Brazil. The authors observed no association between ambient levels of 

O3 and neonatal mortality. 

7.4.9.4. Pos tne o na ta l Morta lity, 1 Mon th to 1 Year 

31 A number of studies focused on the postneonatal period when examining the effects of O3 on 

32 infant mortality. Ritz et al. (2006, 089819) linked birth and death certificates for infants who died 

33 between 1989 and 2000 to evaluate the influence of outdoor air pollution on infant death in the South 

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Coast Air Basin of California. The authors examined exposure periods 2 weeks, 1 month, 2 months, 

and 6 months before each case subject's death and reported no association between ambient levels of 

O3 and postneonatal mortality. Woodruff et al. (2008, 098386) evaluated the county-level 

relationship between cause-specific postneonatal infant mortality and chronic early-life (first 

2 months of life) exposure to air pollutants across the U.S. Similarly, they found no association 

between O3 exposure and deaths from respiratory causes. In the U.K., Hajat et al. (2007, 093276) 

analyzed time-series data of daily infant mortality counts in 10 major cities to quantify any 

associations with short-term changes in air pollution. When the results from the 10 cities were 

combined there was no relationship between O3 and postneonatal mortality, even after restricting the 

analysis to just the summer months. In Ciudad Juarez, Mexico, Romieu et al. (2004, 093074) 

examined the daily number of deaths between 1997 and 2001, estimating the modifying effect of 

SES on the risk of postneonatal mortality. Ambient O3 concentrations were not related to infant 

mortality overall, or in any of the SES groups. In a follow-up study, Carbajal-Arroyo (In Press, 

667773) evaluated the relationship of 1-h daily max O3 levels with postneonatal infant mortality in 

the Mexico City Metropolitan Area between 1997 and 2005. Generally, O3 was not significantly 

related to infant mortality. However, upon estimating the modifying effect of SES on the risk of 

postneonatal mortality, the authors found that O3 was significantly related to respiratory mortality 

among those with low SES. In a separate analysis, the effect of PM10 was evaluated with O3 level 

quartiles. PM10 alone was related to a significant increase in all-cause mortality. The magnitude of 

this effect remained the same when only the days when O3 was in the lowest quartile were included 

in the analyses. However, when only the days when O3 was in the highest quartile were included in 

the analyses, the magnitude of the PM10 effect increased dramatically (OR=1.06 [0.909, 1.241] for 

PM10 on days with O3 in lowest quartile; OR=1.26 [1.08, 1.47] for PM10 on days with O3 in the 

highest quartile. These results suggest that while O3 alone may not have an effect on infant mortality, 

it may serve to potentiate the observed effect of PM10 on infant mortality. 

Tsai et al. (2006, 090709) used a case-crossover analysis to examine the relationship between 

air pollution and postneonatal mortality in Kaohsiung, Taiwan during the period 1994-2000. The risk 

of postneonatal deaths was 1.023 (95% CI: 0.564, 1.858) per 10-ppb increase in 24-h avg O3. The 

confidence interval for this effect estimate is very wide, likely due to the small number of infants that 

died each day, making it difficult to interpret this result. Several other studies conducted in Asia did 

not find any association between O3 concentrations and infant mortality in the postneonatal period. 

Ha et al. (2003, 042552) conducted a daily time-series study in Seoul, Korea to evaluate the effect of 

short-term changes in ambient 8-h O3 concentrations on postneonatal mortality. Son et al. (2008, 

190323) examined the relationship between air pollution and postneonatal mortality from all causes 

among firstborn infants in Seoul, Korea during 1999-2003. Yang et al. (2006, 090760) used a casecrossover analysis to examine the relationship between air pollution exposure and postneonatal 

mortality in Taipei, Taiwan for the period 1994-2000. The authors observed no associations between 

ambient levels of O3 and postneonatal mortality. 

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7.4.9.5. Sud de n Infa nt De a th Syn drome 

The strongest evidence for an association between ambient O3 concentrations and SIDS comes 

from a study that evaluated the county-level relationship between SIDS and chronic early-life (first 

2 months of life) exposure to air pollutants across the U.S.(Woodruff et al., 2008, 098386). The 

authors observed a 1.20 (95% CI: 1.09, 1.32) odds ratio for a 10-ppb increase in O3 and deaths from 

SIDS. There was a monotonic increase in odds of SIDS for each quartile of O3 exposure compared 

with the lowest quartile (highest quartile OR = 1.51; [95% CI: 1.17, 1.96]). In a multi-pollutant 

model including PM10 or PM2.5, CO and SO2, the OR for SIDS and O3 was not substantially lower 

than that found in the single-pollutant model. When examined by season, the relationship between 

SIDS deaths and O3 was generally consistent across seasons with a slight increase for those babies 

born in the summer. When stratified by birth weight, the OR for LBW babies was 1.27 (95% CI: 

0.95, 1.69) per 10-ppb increase in O3 and the OR for normal weight babies was 1.16 (95% CI: 1.01, 

1.32) per 10-ppb increase in O3. 

Conversely, two additional studies reported no association between ambient levels of O3 and 

SIDS. Ritz et al. (2006, 089819) linked birth and death certificates for infants who died between 

1989 and 2000 to evaluate the influence of outdoor air pollution on infant death in the South Coast 

Air Basin of California. The authors examined exposure periods 2 weeks, 1 month, 2 months, and 

6 months before each case subject's death and reported no association between ambient levels of O3 

and SIDS. Dales et al. (2004, 087342) used time-series analyses to compare the daily mortality rates 

for SIDS and the daily air pollution concentrations in 12 Canadian cities during the period of 

1984-1999. Increased daily rates of SIDS were associated with previous day increases in the levels 

of SO2, NO2, and CO, but not O3 or PM2.5. 

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Table 7-6. Brief summary of infant mortality studies 

Study Location Mean O3 (ppb) Exposure 

Assessment Effect Estimate (95% CI): 

Pereira et al. (1998, 

007264) Sao Paulo, Brazil 1-h max: 33.8 Citywide avg L0-2: 1.00 (0.99, 1.01) 

Diaz et al. (2004, 

089894) Madrid, Spain 24-h avg: 11.4 Citywide avg NR 

L0: 0.99 (0.97, 1.02) 

L1: 0.99 (0.96, 1.01) 

Loomis et al. (1999, 

087288) 

Mexico City, 

Mexico 

24-h avg: 44.1 

1-h max: 163.5 1 monitor 

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) 

2 wk before death: 1.03 (0.93, 1.14) 

Ritz et al. (2006, 

089819) southern California 24-h avg: 21.9-22.1 Nearest Monitor 1 mo before death: NR 

2 mo before death: 0.93 (0.89, 0.97) 

6 mo before death: NR 

Hajat et al. (2007, 

093276) 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. (2004, 

095787) Sao Paulo, Brazil 24-h avg: 38.06 Citywide avg L0: 1.00 (0.99, 1.01) 

Ha et al. (2003, 

042552) 

Seoul, South 

Korea 8-h avg: 21.2 Citywide avg L0: 0.93 (0.90, 0.96) 

L1: 0.96 (0.90, 1.03) 

Romieu et al. (2004, 

093074) 

Ciudad Juarez, 

Mexico 8-h avg: 43.43-55.12 Citywide avg 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) 

L0: 1.00 (0.99, 1.00) 

Carbajal-Arroyo et al. 

(In Press, 667773) 

Mexico City, 

Mexico 1-h max: 103.0 Citywide avg 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, 

190323) 

Seoul, South 

Korea 8-ha avg: 25.61 Citywide avg L(NR): 0.984 (0.976, 0.992)* 

Tsai et al. (2006, 

090709) 

Kaohsiung, 

Taiwan 24-h avg: 23.60 Citywide avg L0-2 cum: 1.02 (0.56, 1.86) 

Woodruff et al. (2008, 

098386) 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, 

090760) Taipei, Taiwan 24-h avg: 18.14 Citywide avg L0-2 cum:1.00 (0.62, 1.61) 

L0: NR 

L1: NR 

Dales et al. (2004, 

087342) 12 Canadian cities 24-h: 31.77 Citywide avg 

L2: NR 

L3: NR 

L4: NR 

L5: NR 

Multiday lags of 2-6 days: NR 

#Relative risk of infant mortality per 10 ppb change in O3 

* 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 

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7.4.10. Su mma ry an d Cau s al De te rmina tio n 

The 2006 O3 AQCD concluded that the limited number of studies that investigated O3 

demonstrated no associations between O3 and birth outcomes, with the possible exception of birth 

defects. The current review included an expanded body of evidence on the associations between O3 

and reproductive and developmental effects. Recent epidemiologic and toxicological studies provide 

evidence for an effect of prenatal exposure to O3 on pulmonary structure and function, including 

lung function changes in the newborn, incident asthma, ultrastructural changes in bronchiole 

development, alterations in placental and pup cytokines, and increased pup airway hyper-reactivity. 

Also, there is limited toxicological evidence for an effect of prenatal and early life exposure on 

central nervous system effects, including laterality, brain morphology, neurobehavioral 

abnormalities, and sleep aberration. Recent epidemiologic studies have begun to explore the effects 

of O3 on sperm quality, and provide limited evidence for decrements in sperm concentration, while 

there is limited toxicological evidence for testicular degeneration associated with O3. 

While the collective evidence for many of the birth outcomes examined is generally 

inconsistent (including birth defects), there are several well-designed, well-conducted studies that 

indicate an association between O3 and adverse outcomes. For example, as part of the southern 

California Children's Health Study, Salam et al. (2005, 087885) observed a concentration-response 

relationship of decreasing birth weight with increasing O3 concentrations averaged over the entire 

pregnancy that was clearest above the 30-ppb level (see Figure 7-4). Similarly, Hansen et al. (2008, 

190273) utilized fetal ultrasonic measurements and found a change in ultrasound measurements 

associated with O3 during days 31-60 of gestation indicated that increasing O3 concentration 

decreased an ultrasound measurement for women living within 2 km of the monitoring site. 

There is no evidence that prenatal or early life O3 concentrations are associated with infant 

mortality. Collectively, there is limited though positive toxicological evidence for O3-induced 

developmental effects, including effects on pulmonary structure and function and central nervous 

system effects. Limited epidemiologic evidence for an effect on prenatal O3 exposure on respiratory 

development provides coherence with the effects observed in toxicological studies. There is also 

limited epidemiologic evidence for an association with O3 concentration and decreased sperm 

concentration. A recent toxicological study provides limited evidence for a possible biological 

mechanism (histopathology showing impaired spermatogenesis) for such an association. 

Additionally, though the evidence for an association between O3 concentrations and adverse birth 

outcomes is generally inconsistent, there are several influential studies that indicate an association 

with reduced birth weight and restricted fetal growth. Taking into consideration the positive evidence 

for developmental and reproductive outcomes from toxicological and epidemiological studies, and 

the few influential birth outcome studies, the evidence is suggestive of a causal relationship between 

long-term exposures to O3 and reproductive and developmental effects. 

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7.5. Central Nervous Sys tem Effects 

7.5.1. Effects on the Brain and Behavior 

An epidemiologic study has recently been published examining the association between O3 

exposure and neurobehavioral effects. Chen et al. (2009, 179945) utilized data from the NHANES 

III cohort to study the relationship between O3 levels (mean annual O3 concentration 26.5 ppb) and 

neurobehavioral effects among adults aged 20-59 years. The authors observed an association 

between annual exposure to O3 and tests measuring coding ability (symbol-digit substitution test) 

and attention/short-term memory (serial-digit learning test). Each 10-ppb increase in annual O3 

levels corresponded to an aging-related cognitive performance decline of 3.5 yr for coding ability 

and 5.3 years for attention/short-term memory. These associations persisted in both crude and 

adjusted models. There was no association between O3 levels and reaction time tests. The authors 

conclude that overall, there is an association between long-term O3 exposure and reduced 

performance on neurobehavioral tests. 

In a subchronic study, rats were exposed to 0.25 ppm O3 for 4 h/day for 15-90 days (RivasArancibia et al., 2010, 201544). The exposures caused a complex array of responses, including a 

time-dependent increase in lipid peroxidation products and immunohistochemical changes in the 

hippocampus, that were correlated with decrements in passive avoidance behavioral tests. 

A protective effect of estradiol has been observed in ovariectomized female rats exposed to 

0.25 ppm O3 (4 h/day) for 30 or 60 days (Guevara-Guzmán et al., 2009, 596385). In the olfactory 

bulb, lipid peroxidation was significantly less in rats exposed to O3 and treated daily with estradiol. 

This protective effect of estradiol was also demonstrated for O3-induced decrements in a selective 

olfactory recognition memory test and an olfactory-dependent reward test. Similarly, estradiol 

protected against O3-induced changes in nigral cell morphology and loss of dopamine neurons in rats 

exposed to O3 for 30 days (Angoa-Pérez et al., 2006, 596366). Thus, repeated exposure of rats to O3 

produces lipid peroxidation at multiple sites in the brain and this oxidative stress is accompanied by 

gene expression changes and decrements in behavioral tests. Olfactory changes and loss of 

substantia nigra neurons are associated with Parkinson's disease in humans. Inhibition of these 

effects with estradiol treatment is consistent with the higher incidence of Parkinson’s disease in men 

and the amelioration of Parkinsonian symptoms by estrogen therapy. 

Adverse CNS effects have also been demonstrated in newborn and adult rats whose only 

exposure to O3 occurred in utero. Several neurotransmitters were assessed in male offspring of dams 

exposed to 1 ppm O3 during the entire pregnancy (Gonzalez-Pina et al., 2008, 475317). The data 

showed that catecholamine neurotransmitters were affected to a greater degree than indole-amine 

neurotransmitters in the cerebellum. Adverse CNS changes, including behavioral, cellular, and 

biochemical effects, have also been observed after in utero exposure to 0.5 ppm O3 for 12 h/day from 

GD5 to GD20 (Boussouar et al., 2009, 596368). Tyrosine hydroxylase labeling in the nucleus tractus 

solitarius was increased after in utero exposure to O3 whereas Fos protein labeling did not change. 

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1 When these offspring were challenged by immobilization stress, neuroplasticity pathways, which 

2 were activated in air offspring, were inhibited in O3 offspring. Although the effect of O3 exposure 

3 concentration was not studied in these two in utero studies, it has been examined in one study. 

4 Santucci et al. (2006, 596414) investigated behavioral effects and gene expression after in utero 

5 exposure of mice to as little as 0.3 ppm O3. Increased defensive/submissive behavior and reduced 

6 social investigation were observed in both the 0.3- and 0.6-ppm O3 groups. Changes in gene 

7 expression of brain-derived neurotrophic factor (BDNF, increased in striatum) and nerve growth 

8 factor (NGF, decreased in hippocampus) accompanied these behavioral changes. Thus, these three 

9 studies demonstrate that CNS effects can occur as a result of in utero exposure to O3, and although 

10 the mode of action of these effects is not known, it has been suggested that circulating lipid 

11 peroxidation products may play a role (Boussouar et al., 2009, 596368). Importantly, these adverse 

12 CNS effects occurred in rodent models after in utero only exposure to (semi-) relevant 

13 concentrations of O3. 

14 

7.5.2. Su mma ry an d Cau s al De te rmina tio n 

15 The 2006 O3 AQCD (U.S. EPA, 2006, 088089) included toxicological evidence that acute 

16 exposures to O3 are associated with alterations in neurotransmitters, motor activity, short and long 

17 term memory, and sleep patterns. Additionally, histological signs of neurodegeneration have been 

18 observed. However, evidence regarding chronic exposure and neurobehavioral effects was not 

19 available. Recent research in the area of O3-induced neurotoxicity has included several long-term 

20 exposure studies. Notably, the first epidemiologic study to examine the relationship between O3 

21 exposure and neurobehavioral effects observed an association between annual O3 levels and an 

22 aging-related cognitive performance decline in tests measuring coding ability and attention/short23 term memory. This observation is supported by studies in rodents which demonstrate oxidative stress 

24 in the brain and associated decrements in behavioral tests, including those measuring memory, after 

25 subchronic exposure to 0.25 ppm O3. Additionally, neurobehavioral changes are evident in animals 

26 whose only exposure to O3 occurred in utero. Collectively, the limited epidemiologic and 

27 toxicological evidence is coherent and suggestive of a causal relationship between O3 exposure 

28 and adverse CNS effects. 

7.6. Carcinogenic and Genotoxic Potential of Ozone 

7.6.1. Introd uc tion 

29 The radiomimetic and clastogenic qualities of O3, combined with its ability to stimulate 

30 proliferation of cells in the respiratory tract, have suggested that O3 could act as a carcinogen. 

31 However, toxicological studies of tumorigenesis in the rodent lung have yielded mixed and often 

32 confusing results, and the epidemiologic evidence is equally conflicted. The 2006 O3 AQCD 

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1 concluded that, “the weight of evidence from recent animal toxicological studies and a very limited 

2 number of epidemiologic studies do not support ambient O3 as a pulmonary carcinogen” (U.S. EPA,
 

3 2006, 088089). 

4 Multiple epidemiologic studies reported in the 2006 O3 AQCD examined the direct association 

between O3 exposure and cancer. The largest of these studies, by Pope et al. (2002, 024689), 

6 included 500,000 adults from the American Cancer Society’s (ACS) Cancer Prevention II study. In 

7 this study, no association was observed between O3 and lung cancer mortality. The Adventist Health 

8 Study of Smog (AHSMOG) also examined the association between O3 and lung cancer mortality 

9 (Abbey et al., 1999, 047559). There was a positive association between O3 levels and lung cancer 

mortality among men. No association was reported for women. Another study using the AHSMOG 

11 cohort assessed the risk of incident lung cancer (Beeson et al., 1998, 048890). Among males, an 

12 association with incidence of lung cancer was observed with increasing O3 concentrations. When 

13 stratified by smoking status, the association persisted among never smokers but was null for former 

14 smokers. No association was detected for females. The Six Cities Study examined various air 

pollutants and mortality but did not specifically explore the association between O3 concentrations 

16 and lung cancer mortality due to low variability in O3 levels across the cities (Dockery et al., 1993, 

17 044457). An ecologic study performed in Sao Paulo City, Brazil examined the correlations between 

18 O3 levels in four of the city districts and incident cancer of the larynx and lung reported in 1997 

19 (Pereira et al., 2005, 073851). A correlation between the average number of days O3 levels exceeded 

air quality standards from 1981 to 1990 and cancer incidence was present for larynx cancer but not 

21 for lung cancer. 

22 Early toxicological research demonstrated lung adenoma 1 acceleration in mice with daily 

23 exposure to 1 ppm over 15 months (Stokinger, 1962, 015101). Later work demonstrated a significant 

24 increase in lung tumor numbers in one strain of mouse (A/J) but not another after exposure to 

0.3-0.8 ppm O3 (Hassett et al., 1985, 040704; Last et al., 1987, 040830). The A/J mouse strain is 

26 known to have a high incidence of spontaneous adenomas, and further studies using this strain found 

27 a statistically significant increase in lung tumor incidence after a 9-month exposure to 0.5 ppm and 

28 incidence and multiplicity after a 5 month exposure to 0.12 ppm with a 4-month recovery period 

29 (Witschi et al., 1999, 011602). However, these findings were discounted by the study authors due to 

the lack of a clear dose response, and results from the Hassett et al. and Last et al. studies were 

31 retrospectively deemed spurious based on what appeared to be unusually low spontaneous tumor 

32 incidences in the control groups (Witschi, 1991, 042509). A study of carcinogenicity of O3 by the 

33 National Toxicology Program (NTP, 1994, 011143) reported increased incidences of 

34 alveolar/bronchiolar adenoma or carcinoma (combined) in female B6C3F1 mice exposed over 

2 years or a lifetime to 1.0 ppm and marginally increased incidences in male mice exposed to 0.5 and 

36 1.0 ppm. Thus there was equivocal evidence of carcinogenic activity in male mice and some 

1 NOTE: Although adenomas are benign, over time they may progress to become malignant, at which point they are called 

adenocarcinomas. Adenocarcinoma is the predominant lung cancer subtype in most countries, and is the only lung cancer found in 

nonsmokers. From page 8-33 of the 1970 O3 AQCD: “No true lung cancers have been reported, however, from experimental exposures 

to either O3 alone or any other combination or ingredient of photochemical oxidants.” 

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1 evidence of carcinogenic activity of O3 in females. Some semblance of a dose-response relationship 

2 was also evident in this study. 

3 In Fischer-344/N rats (50 of each sex per group), neither a 2-year nor lifetime exposure to O3 

4 ranging from 0.12 to 1.0 ppm was found to be carcinogenic (Boorman et al., 1994, 038195). 

However, a marginally significant carcinogenic effect of 0.2 ppm O3 was reported in a study of male 

6 Sprague-Dawley rats exposed for 6 months (n = 50) (Monchaux et al., 1996, 086197). These two 

7 studies also examined co-carcinogenicity of O3 with NNK1 (Boorman et al., 1994, 038195) or a 

8 relatively high dose of radon (Monchaux et al., 1996, 086197), finding no enhancement of NNK 

9 related tumors and a slight non-significant increase in tumor incidence after combined exposure with 

radon, respectively. Another study exploring co-carcinogenicity was conducted in hamsters. Not only 

11 was there no enhancement of chemically induced tumors in the peripheral lung or nasal cavity, but 

12 results suggested that O3 could potentially delay or inhibit tumor development (Witschi et al., 1993, 

13 043206). Thus there is no concrete evidence that O3 can act as a co-carcinogen. 

14 Immune surveillance is an important defense against cancer, and it should be noted that natural 

killer (NK) cells, which destroy tumor cells in the lung, appear to be inhibited by higher doses of O3 

16 and either unaffected or stimulated at lower doses (Section 6.2.5.4, Infection and Adaptive 

17 Immunity). This aspect of tumorigenesis adds yet another layer of complexity which may be 

18 reflected by conflicting results across studies. 

19 The following sections will examine epidemiologic studies of cancer incidence and mortality 

that have been published since the 2006 O3 AQCD. One study has been published with cancer as the 

21 outcome; most studies examine markers of exposure or susceptibility and will be reported on later in 

22 this section. Recent toxicological studies are also described. 

7.6.2. Lu ng Ca n ce r Incid e nc e a nd Mo rtality 

23 A recent re-analysis of the full ACS CPSII cohort by the Health Effects Institute is the only 

24 epidemiologic study that has explored the association between O3 and cancer mortality since the last 

O3 AQCD. Krewski et al. (2009, 191193) conducted an extended follow-up of the cohort 

26 (1982-2000). Mean O3 levels [obtained from the Aerometric Information Retrieval System (AIRS) 

27 for 1980] were 22.91 ppb for the full year and 30.15 ppb for the summer months (April-September). 

28 No association was reported between lung cancer mortality and O3 (HR 1.00 [95% CI: 0.96-1.04] 

29 per 10 ppb O3). Additionally, no association was observed when O3 was restricted to the summer 

months. There was also no association present in a sub-analysis of the cohort examining the 

31 relationship between O3 and lung cancer mortality in the Los Angeles area. 

32 Since the 2006 O3 AQCD, two toxicological studies have examined potential carcinogenicity 

33 of O3 (Kim and Cho, 2009, 200775; Kim and Cho, 2009, 200773). Looking across both studies, 

34 which used the same mouse strain as the National Toxicology Program study described above, 

0.5 ppm O3 alone or in conjunction with chemical tumor inducers did not enhance lung tumor 

1 4-(N-nitrosomethylamino)-1-(3-pyridyl)-1-butanone 

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1 incidence in males or females. However, a 10% incidence of oviductal carcinoma was observed in 

2 mice exposed to 0.5 ppm O3 for 16 weeks. The implications of this observation are unclear, 

3 particularly in light of the lack of statistical information reported. Additionally, there is no mention 

4 of oviductal carcinoma after 32 weeks of exposure, and no oviductal carcinoma was observed after 

one year of exposure. 

7.6.3. DNA Damage 

6 The potential for genotoxic effects relating to O3 exposure was predicted from the 

7 radiomimetic properties of O3. The decomposition of O3 in water produces OH and HO2 radicals, the 

8 same species that are generally considered to be the biologically active products of ionizing 

9 radiation. Ozone has been observed to cause degradation of DNA in a number of different models 

and bacterial strains. Until the reports of Zelac et al.(1971, 039743; 1971, 039752), the toxic effects 

11 of O3 were generally assumed to be confined to the tissues directly in contact with the gas, such as 

12 the respiratory epithelium. Due to the highly reactive nature of O3, little systemic absorption was 

13 predicted. Zelac et al., however, reported a significant increase in chromosome aberrations in 

14 peripheral blood lymphocytes from Chinese hamsters exposed to 0.2 ppm for 5 hours. Other in vivo 

exposure studies found increased DNA strand breaks in respiratory cells from guinea pigs (Ferng et 

16 al., 1997, 083578) and mice (Bornholdt et al., 2002, 036677) but only with exposure to higher doses 

17 of O3 (1 ppm for 72 hours and 1 or 2 ppm for 90 minutes, respectively). In other studies there were 

18 no observations of chromosomal aberrations in germ cells, but mutagenic effects have been seen in 

19 offspring of mice exposed to 0.2 ppm during gestation (blepharophimosis or dysplasia of the 

eyelids). The overall evidence for mutagenic activity from in vitro studies is positive, and in the 

21 National Toxicology Program report described above, O3 was found to be mutagenic in Salmonella, 

22 with and without S9 metabolic activation. No new toxicological studies of DNA damage have 

23 become available since the 2006 O3 AQCD. 

24 A number of epidemiologic studies looked at the association between O3 and DNA and cellular 

level damages. These changes may be relevant to mechanisms leading to cancers development and 

26 serve as early indicators of elevated risk of mutagenicity. 

27 Two studies performed in California examined cytogenetic damage in relation to O3 

28 exposures. Huen et al. (2006, 089035) examined cytogenetic damage among African American 

29 children and their mothers in Oakland, CA. Increased O3 (mean monthly 8-h O3 concentrations 

ranged from about 30 ppb in April to 14 ppb in November) was associated with increased 

31 cytogenetic damage (micronuclei frequency among lymphocytes and buccal cells) even after 

32 adjustment for household/personal smoking status and distance-weighted traffic density. Chen et al. 

33 (2006, 196504) recruited college students at the University or California, Berkeley who reported 

34 never smoking and compared their levels of cytogenetic damage (micronuclei frequency from buccal 

cells) in the spring and fall. Cytogenetic damage was greater in the fall, which the authors attributed 

36 to the increase in O3 over the summer. However, O3 levels over 2, 7, 10, 14, or 30 days 

37 (concentrations not given) before collection of buccal cells did not correlate with cytogenetic 

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damage. Estimated lifetime O3 exposure was also not correlated with cytogenetic damage. 

Additionally, the authors exposed a subset of the students (n=15) to 200 ppb O3 for 4 hours while the 

students exercised intermittently. Ozone was found to be associated with an increase in cytogenetic 

damage in degenerated cells but not in normal cells 9-10 days after exposure. Increased cytogenetic 

damage was also noted in peripheral blood lymphocytes collected 18 hours after exposure. 

A study performed in Mexico recruited 55 male workers working indoors (n=27) or outdoors 

(n=28) in Mexico City or Puebla, Mexico in order to study the relationship between O3 and DNA 

damage (detected from peripheral blood samples using the Comet assay) (Tovalin et al., 2006, 

091322). The median estimated daily O3 concentrations were estimated to be 28.5 ppb for outdoor 

workers and 5.1 ppb for indoor workers in Mexico City and 36.1 ppb for outdoor workers and 

19.5 ppb for indoor workers in Puebla. Overall, a positive correlation between O3 levels and DNA 

damage was observed. However, when examining the relationship by city and workplace, only DNA 

damage in outdoor workers in Mexico City remained correlated with O3 levels. 

Three studies examining the relationship between O3 and DNA-level damage have been 

performed in Europe. The largest of these was the GenAir case-control study, which was nested 

within the European Prospective Investigation into Cancer and Nutrition (EPIC) study, and included 

individuals recruited between 1993 and 1998 from ten European countries. Only non-smokers (must 

not have smoked for at least 10 years prior to enrollment) were enrolled in the study. The researchers 

examined DNA adduct levels (DNA bonded to cancer-causing chemicals) and their relationship with 

O3 concentrations (concentrations not given) (Peluso M Hainaut et al., 2005, 089057). A positive 

association was seen between DNA adduct levels and O3 concentrations from 1990-1994 but not O3 

levels from 1995-1999. In adjusted analyses with DNA adduct levels dichotomized as high and low 

(detectable versus non-detectable), the OR was 1.97 (95% CI: 1.08, 3.58) when comparing the upper 

tertile of O3 concentration to the lower two tertiles. Two other European studies were conducted in 

Florence, Italy. The most recent of these enrolled individuals from the EPIC study into a separate 

study between March and September of 1999 (Palli et al., 2009, 196688). The purpose of the study 

was to examine oxidative DNA damage (determined by Comet assay using blood lymphocytes) in 

association with varying periods of O3 exposure. The researchers observed that longer periods of 

high O3 exposure (concentrations not given) were more strongly correlated with oxidative DNA 

damage than shorter exposures (i.e., the rho [p-value] was 0.26 [0.03] for 0-10 days and 0.35 [0.002] 

for 0-90 days). This correlation was stronger among men compared to women. The correlations for 

all time periods had p-values <0.05 for ex- and never-smokers. For current smokers, the correlation 

was only observed among time periods≤ 25 days. When adjusted for age, gender, smoking history, 

traffic pollution exposure, period of blood draw, and area of residence, the association between O3 

levels and oxidative DNA damage was positive for O3 levels 0-60 days, 0-75 days, and 0-90 days 

prior to blood draw. Positive, statistically significant associations were not observed among shorter 

time periods. The other study performed in Florence recruited healthy volunteers who reported being 

non-smokers or light smokers (Giovannelli et al., 2006, 199894). The estimated O3 levels during the 

study ranged from approximately 4-40 ppb for 3-day averages, 5-35 ppb for 7-day averages, and 

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1 7.5-32.5 ppb for 30-day averages. Ozone concentrations were correlated with DNA strand breaks
 

2 (measured from blood lymphocytes) over longer exposure periods (p-value: 0.002 at 30 days, 

3 p-value: 0.04 at 7 days; p-value: 0.17 at 3 days). This association was robust to control for
 

4 temperature, solar radiation, gender, and age. No association was seen between O3 concentrations
 

and measures of oxidative DNA damage at 3, 7, or 30 days.
 

7.6.4. Su mma ry an d Cau s al De te rmina tio n 

6 The 2006 O3 AQCD (U.S. EPA, 2006, 088089) reported that evidence did not support ambient 

7 O3 as a pulmonary carcinogen. Since the 2006 O3 AQCD, very few epidemiologic and toxicological 

8 studies have been published that examine O3 as a carcinogen, but collectively, study results indicate 

9 that O3 may contribute to DNA damage. Overall, the evidence is inadequate to determine if a 

causal relationship exists between ambient O3 exposures and cancer. 

7.7. Mortality 

11 A limited number of epidemiologic studies have assessed the relationship between long-term 

12 exposure to O3 and mortality in adults. The 2006 O3 AQCD concluded that an insufficient amount of 

13 evidence existed “to suggest a causal relationship between chronic O3 exposure and increased risk 

14 for mortality in humans” (U.S. EPA, 2006, 088089). In addition to the infant mortality studies 

discussed in Section 7.4.9, two additional studies have been conducted among adults since the last 

16 review; an ecologic study that finds no association between mortality and O3, and a reanalysis of the 

17 ACS cohort that specifically points to a relationship between long-term O3 exposure and an 

18 increased risk of respiratory mortality. These studies supplement the evidence from long-term cohort 

19 studies characterized in previous reviews of O3, (U.S. EPA, 1996, 017831; U.S. EPA, 2006, 088089) 

and are summarized here briefly. 

21 In the Harvard Six Cities Study (Dockery et al., 1993, 044457), adjusted mortality rate ratios 

22 were examined in relation to long-term mean O3 concentrations in six cities: Topeka, KS; St. Louis, 

23 MO; Portage, WI; Harriman, TN; Steubenville, OH; and Watertown, MA. Mean O3 concentrations 

24 from 1977 to 1985 ranged from 19.7 ppb in Watertown to 28.0 ppb in Portage. Long-term mean O3 

concentrations were not found to be associated with mortality in the six cities. However, the authors 

26 noted that “The small differences in O3 levels among the (six) cities limited the power of the study to 

27 detect associations between mortality and O3 levels.” In addition, while total and cardio-pulmonary 

28 mortality were considered in this study, respiratory mortality was not specifically considered. 

29 In a subsequent large prospective cohort study of approximately 500,000 U.S. adults, Pope et 

al. (2002, 024689) examined the effects of long-term exposure to air pollutants on mortality 

31 (American Cancer Society, Cancer Prevention Study II). All-cause, cardiopulmonary, lung cancer 

32 and other mortality risk estimates for long-term O3 exposure are shown in Figure 7-5. While no 

33 consistently significant positive associations were observed between O3 and mortality, the mortality 

34 risk estimates were larger when analyses considered more accurate exposure metrics, rising when the 

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1 entire period was considered (versus just at the start of the study) and becoming marginally 

2 significant when the exposure estimate was restricted to the summer months (July to September), 

3 especially when considering cardio-pulmonary deaths. In contrast, consistent positive and significant 

4 effects of PM2.5 were observed for both lung cancer and cardio-pulmonary mortality. 

Years of Data Number of Number of Participants 1-h Max O3 Mean (SD) Collection Metropolitan Areas (in thousands) 

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: Derived with permission from American Medical Association, Pope et al. (2002, 024689). 

Figure 7-5. Adjusted ozone-mortality relative risk estimates (95% CI) by time period of analysis 

per subject-weighted mean ozone concentration in the Cancer Prevention Study II

by the American Cancer Society. 

5 A study by Abbey et al. (1999, 047559) examined the effects of long-term air pollution 

6 exposure, including O3, on all-cause (n = 1,575), cardiopulmonary (n = 1,029), nonmalignant 

7 respiratory (n = 410), and lung cancer (n = 30) mortality in the long-term prospective Adventist 

8 Health Study of Smog (AHSMOG) of 6,338 nonsmoking, non-Hispanic white individuals living in 

9 California. A particular strength of this study was the extensive effort devoted to assessing long-term 

10 air pollution exposures, including interpolation to residential and work locations from monitoring 

11 sites over time and space. No associations with long-term O3 exposure were observed for all cause, 

12 cardiopulmonary, and nonmalignant respiratory mortality. In a follow-up, Chen et al. (2005, 087942) 

13 utilized data from the AHSMOG study and reported no significant associations between long-term 

14 O3 exposure (mean O3 concentration 26.2 ppb) and fatal coronary heart disease. Thus, no association 

15 of chronic O3 exposure with mortality outcomes has been detected in this study. 

16 Lipfert et al. (2000, 004087; 2003, 052250) reported positive effects on all-cause mortality for
 

17 peak O3 exposures (95th percentile levels) in the U.S. Veterans Cohort study of approximately 

18 50,000 middle-aged men recruited with a diagnosis of hypertension. The actual analysis involved 

19 smaller subcohorts based on exposure and mortality follow-up periods. Four separate exposure 

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periods were associated with three mortality follow-up periods. For concurrent exposure periods, 

peak O3 was positively associated with all-cause mortality, with a 9.4% (95% CI: 0.4, 18.4) excess 

risk per mean 95th percentile O3 less estimated background level (not stated). “Peak” refers, in this 

case, to the 95th percentile of the hourly measurements, averaged by year and county. In a further 

analysis, Lipfert et al. (2003, 052250) reported the strongest positive association for concurrent 

exposure to peak O3 for the subset of subjects with low diastolic blood pressure during the 1982 to 

1988 period. Two more recent studies of this cohort focused specifically on traffic density (Lipfert et 

al., 2006, 088756; 2006, 088218). Lipfert (2006, 088218) concluded that: “Traffic density is seen to 

be a significant and robust predictor of survival in this cohort, more so than ambient air quality, with 

the possible exception of O3,” reporting a significant O3 effect even with traffic density included in 

the model: RR=1.080 per 40 ppb peak O3 (95% CI: 1.019, 1.146). In Lipfert (2006, 088756), which 

considers only the EPA Speciation Trends Network (STN) sites, O3 drops to non-significant predictor 

of total mortality for this cohort. However, the authors acknowledge that: “Peak O3 has been 

important in analyses of this cohort for previous periods, but in the STN data set, this variable has 

limited range and somewhat lower values and its small coefficient of variation results in a relatively 

large standard error.” The restriction to subjects near STN sites likely reduced the power of this 

analysis, though the size of the remaining subjects considered was not reported in this paper. In 

addition, these various Veteran's Cohort studies considered only total mortality, and did not consider 

mortality on a by-cause basis. 

An ecological study in Brisbane, Australia used a geospatial approach to analyze the 

association of long-term exposure to gaseous air pollution with cardio-respiratory mortality, in the 

period 1996-2004 (Wang et al., 2009, 199990). A generalized estimating equations model was 

employed to investigate the impact of NO2, O3 and SO2, but PM was not addressed. The results 

indicated that long-term exposure to SO2 was associated with cardio-respiratory mortality, but the 

fact that this study considered only one city, and that the range of O3 exposure across that city 

(23.7-35.6 ppb) was low and slight in variation in comparison to the range of other pollutants across 

the city, limited study power. In addition, confounding factors (e.g., smoking) could not be addressed 

at the individual level in this ecological study. Respiratory mortality was not evaluated separately. 

In the most recent follow-up analysis of the ACS cohort (Jerrett et al., 2009, 194160), 

cardiopulmonary deaths were subdivided into respiratory and cardiovascular, separately, as opposed 

to combined in the Pope et al. (2002, 024689) work. This analysis utilized the ACS cohort with data 

from 1977 through 2000 (mean O3 concentration ranged from 33.3 to 104.0 ppb). A 10-ppb 

increment in exposure to O3 elevated the risk of death from the cardiopulmonary, cardiovascular, 

ischemic heart disease, and respiratory causes. Inclusion of the concentration of PM2.5 measured in 

1999-2000 as a co-pollutant attenuated the association with exposure to O3 for all end points except 

death from respiratory causes, for which a significant association persisted (Table 7.7). The 

association between increased O3 concentrations and increased risk of death from respiratory causes 

was insensitive to the use of a random-effects survival model allowing for spatial clustering within 

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

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1 considered individually. Subgroup analyses showed that temperature and region of country, but not 

2 sex, age at enrollment, body-mass index, education, or PM2.5 concentration, modified the effects of 

3 O3 on the risk of death from respiratory causes (i.e., risks were higher at higher temperature, and in 

4 the Southeast, Southwest, and Upper Midwest). Ozone threshold analyses indicated that the 

5 threshold model was not a better fit to the data (p > 0.05) than a linear representation of the overall 

6 O3-mortality association. Overall, this new analysis indicates that long-term exposure to PM2.5 

7 increases risk of cardiac death, while long-term exposure to O3 is specifically associated with an 

8 increased risk of respiratory death. 

Table 7-7. Relative risk (and 95% CI) of death attributable to a 10-ppb change in the ambient ozone

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) 

Source: Used with permission from Massachusetts Medical Society, Jerrett et al. (2009, 194160) 

7.7.1. Su mma ry an d Cau s al De te rmina tio n 

9 The 2006 O3 AQCD concluded that an insufficient amount of evidence existed “to suggest a 

10 causal relationship between chronic O3 exposure and increased risk for mortality in humans” 

11 (U.S. EPA, 2006, 088089). Two additional studies have been conducted since the last review, an 

12 ecologic study that finds no association between mortality and O3 (Wang et al., 2009, 199990), and a 

13 re-analysis of the ACS cohort that specifically points to a relationship between long-term O3 

14 exposure and an increased risk of respiratory mortality (Jerrett et al., 2009, 194160). The findings 

15 from the Jerrett et al. (2009, 194160) study are consistent and coherent with the evidence from 

16 epidemiologic, controlled human exposure, and animal toxicological studies for the effects of short17 and long-term exposure to O3 on respiratory effects. Additionally, the evidence for short- and long18 term respiratory morbidity provides biological plausibility for mortality due to respiratory disease. 

19 Collectively, the evidence is suggestive of a causal relationship between long-term O3 exposures 

20 and all-cause mortality. 

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