# 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 O 3 . Both epidemiologic and toxicological studies provide a basis



### 3 for examining long

- term O 3 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 O 3 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 e ach 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, repro ductive/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. Determ ination 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 O 3 AQCD (U.S. EPA, 2006, 088089 ) examined evidence for



### 18 relationships between long

- term O 3 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. T he term seasonal was used in these studies as a measure of a long

- term exposure of



### 21 several months. Animal toxicology data provide d a clearer picture indicating that long

- term O 3 

### 22 exposure may have lasting effects. Chronic exposure studies in animals have re ported biochemical



### 23 and morphological changes suggestive of irreversible long

- term O 3 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 O 3 exposure wa s associated with small er increases in l ung 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 O 3 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, NO 2 , and PM 2.5 . A sthmatic children



### 6 with GSTM1 null genotype were found to be more susceptible to the impact of O 3 exposure on small



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



### 8 examined the relationship between long

- term O 3 exposures and inflammation. Inflammatory effects



### 9 consistent with known effects of O 3 such as increased eosi nophil levels were observed in an Austrian
study (Frischer et al., 2001, 019683 ) . The cross- sectional surveys available fo r the 2006 O 3 AQCD



### 11 detected no associations between long

- term O 3 exposures and asthma prevalence, asthma

- related



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



### 13 N ew evidence presented below reports consistent assoc iations between long

- term O 3 exposure


### 14 and new

- onset asthma related to genotype in U.S. cohorts in multi

- community studies. Related
studies report coherent relationships between resp iratory symptoms among asthma tics and long

- term



### 16 O 3 exposure. A new line of evidence reports a positive exposure response relationship between first



### 17 asthma hospitalization and long

- term O 3 exposure . Results from recent studies examining p ulmonary



### 18 function, inflammation, and allergic response s 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 mechani sms for these complex diseases. Oxidative stress likely underlies these mechanistic hypotheses (Gilliland



### 26 et al., 1999, 155792 ) . Susceptibili ty 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 )

. E vidence for a gene

pollution interacti on 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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39 Evidence for associations between long

- term exposure to O 3 and new

- onset asthma is provided
by new s tudies 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 y ea rs 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 fun ctiona l polymorphisms of HMOX

- 1 [(GT) n repeat] , CAT

(- 262C > T

- 844C > T0, and MNSOD (Ala

- 9Val) ar e associated with new

- onset asthma was

evaluated , and then whether the effects of these variants var ied by exposure to O 3 (Islam et al., 2008,

097348 ). HMOX1 [ heme oxygena se (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 sourc e 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

. T he 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 -O 3 communities.
The subjects were obtained from the CHS from 12 communities in south ern California.

Child ren with a history of asthma or wheeze were excluded from this analysis. Analys e s 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

relation ship 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 diagn osed (Islam et al., 2008,

097348 ) . The evidence indicated that t he effect of variation in the HMOX

- 1 gene on risk of new

onset asthm a differed by ambient O 3 level

. An interaction P value wa s reported of 0.003 from the
hierarchical two stage Cox proportional hazard model fitting the community

- specific O 3 and PM 10
levels (continuous) and controlling for random effect of the communities. The annual O 3 levels

(10 :00 a. m. – 6 :00 p.m .) ranged from 46.5 to 64.9 ppb in the six higher O 3 communities (mean =

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38 55.2 ppb ) and 28.6 to 45.5 ppb in the six lower O 3 communities (mean = 3 8.4 ppb )

. Average O 3 levels showed low correlation with the other monitored pollutants. The interaction indicate d a greater
effect (association) of community O 3 level among children with the gene than with children without
the gene. Alleles with 23 or fewe r (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 larges t
protective effect of the (GT) n repeat polymorphism of H M OX

- 1 was observed for children w ho

were S

- allele carriers and resided in low -O 3 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 O 3 communities (HR 0.88; [ 95%

CI : 0.33, 2.34] ) was twofold greater than in those who r esided in the low -O 3 communities (HR 0.44).

The non

- parallelism of the two lines in Figure 7- 1 illustrates the interaction: Children with the S

allele have protection against the onset of asthma ; however, in high - O 3 communities, this protection

is attenua ted. T he 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 intera ctions

were observed between PM 10 or other pollutants and the H M OX

- 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 th e genes was based on a priori hypotheses defined by a well

- studied biological
pathway. Thus in this cohort in south ern California, Islam et al. (2008, 097348 ) relate d new

- onset

asthma to O 3 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 O 3 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 O 3 specifically or consi der exercise

. A primary conclusion that the authors (Islam et al.,

2009, 196715 ) report ed 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 ne w onset asthma
compared with those without the null genotype. Further, the CHS investigators examine d the
complex interrelationship of antioxidant defenses with asthma risk with increasing doses of O 3 , 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 O 3 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

increa sed 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 mor e team
sports and lived in high -O 3 communities demonstrates the pote ntial importance of a combination of
genetic variability, O 3 exposur e and behavior on asthma risk.

March 20 11 7

- 4 DRAFT – DO NOT CITE OR QUOTEInteraction of Gene presence and Ozone Level on the Hazard Ratio of New Onset Asthma (P-value of 0.003)

0.94 0.88 1

0.44

0 0.5

1 1.5

2 2.5

3

Community Mean Ozone LevelHR

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

comm unity 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 asth ma are limited (Li et al.,



### 2 2006, 596447 ). Asthmatic s ubjects 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 O 3 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 O 3 . Among atopic children living in high O 3 communities, those carrying the ARG1



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



### 12 P heterogeneity across atopy/ O 3 categories = 0.008).



### 13 Further, the CHS presents results examining the relationship of new onset asthma with traffic



### 14 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 schoo ls. The



### 16 HR was 0.76 (95% CI : 0.38, 1.54) across the range of ambient O 3 exposure in the communities. With

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38 adjustment for school and residential non

- freeway traffic

- related exposure, the estimated HR for O 3 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 O 3 AQ CD observed positive relationships
between chronic exposure to O 3 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.

I n 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 f rom elementary and middle schools within 1 km of air monitoring stations. The risk of

asthma was related to O 3 in the one

- pollutant model . T he addition of other pollutants , in two

pollutant and three

- pollutant models

, increased the O 3 risk estimates. The p revalence of childhood
asthma was assessed in Portugal by contrasting the risk of asthma between a high O 3 rural area and

an area with low O 3 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 O 3 levels without the

confoun ding effects of other pollutants. Both evaluation for asthma symptoms and FEV 1 suggested

that O 3 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 y ea rs of age in a popula tion

- 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 us ing 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 . T raffic

- related pollutants were associated with the hig hest risk . Ozone was inversely correlated with the primary traffic

- relate d 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 ) . I n

a cross

- sectional analysis, Akinbamia et al. (2010, 378580 ) examine d the association between
chronic exposure to outdoor pollutants (12

- mo n th average levels by county) and asthma outcomes in
a national sample of children ages 3

- 17 y ea rs living in U.S. metropolitan areas (National Health
Interview Survey, N = 34,073). A 5- ppb increase in estimated 8- h max O 3 concentration (annual

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

ast hma 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

- mo n th

average O 3 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 O 3 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 evid ence relating O 3 exposure and the risk of asthma.

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### 1 T he occurrence of bronchitic symptoms among children with asthma was investigated in the



### 2 CHS examining the role of gene

- en vironment interactions and long

- term O 3 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 O 3 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

- alle le



### 7 (e.g.

, GA or AA). Low

- versus high-O 3 stra ta were defined as less than or greater than 50

- ppb O 3 

### 8 average

. Asthmatic children with TNF

- 308 GG genotype had a significantly reduced risk of



### 9 bronchi ti c symptoms with low -O 3 ex posure (OR: 0.53; [ 95% CI: 0.31, 0.91 ] ). The risk was not



### 10 reduced in children living in high -O 3 comm unities (OR: 1.42; [ 95% CI: 0.75, 2.70] ). The difference



### 11 in genotypic effects between low

- and high -O 3 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 adjuste d O 3 community

- specific beta

- coefficients plotted against ambient O 3 concentration,



### 16 us ing 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 PM 10 , 

### 18 PM 2.5 , NO 2 , acid vapor or second

- hand smoke exposure. T hese results sugg est a role of gene



### 19 environment interactions such as long

- term O 3 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 Using dummy variables in each of the 12 ﷒ children w ith asthma in the CHS. communities

, betas were calculated of TNF

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

.

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39 T he 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., 20 09, 196720 ). In this

cross

- sectional study, a sthma severity over the past 12 mon ths was assessed among 328 adult
asthmatics using two methods: ( 1) a four

- cl ass severity score that integrated clinical events and t ype

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 w h ere a total of 93%
of the subjects lived within 10 km of a monitor

, but w h ere 70% of the O 3 concentrations were back

extrapolated values ; and (2 [second method] ) a validated spatial model that used geostatistical
interpolations and then ass igned air pollutants to the geo code d residential addresses of all
participants and individually assigned exposure to ambient air pollution estimates. Higher asthma
severity score s were significantly related to both the 8- h avg O 3 during April

- September and the
number of days with 8-h O 3 ave rages above 55 ppb . Both exposure assessment methods and severity
score methods resulted in very similar findings. Effect estimates of O 3 were similar in three -pollutant

models. No PM data w ere available

. Since these estimates were not sensitive to the inc lusion of

ambient NO 2 in the three

- pollutant models, the authors viewed the findings not to be explained by
particles which usually have substantial correlations between PM and NO 2 . Ozone concentrations by

the first method for annual levels, nearest monito r were 8- hour s (n = 210) ; mean ( + SD) 30.25 ( +9.7 );

IQR: 21

- 36.5 ( + 15.5) ppb . The second exposure approach yielded summer (n = 308) ; m ean ( + SD)

levels of 31.5 ( +5.2 ); IQR of 28.5

- 33.9 ( +5.5) ppb . Effect estimates for O 3 in three

- pollutant models
including O 3 , SO 2 , and NO 2 yielded OR for O 3 - days of 2.74 (95% CI : 1.68, 4.48) per IQR days of

10

- 28 ( + 18) ppb

. The effect estimates for SO 2 and NO 2 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 ac count duration of residence
did not change the result. This study suggests that a higher asthma severity score is related to long

term O 3 exposure.

T he interrelationships between variants in catalase (CAT) and myeloperoxidase (MPO) genes, ambient polluta nts, and acute respiratory illness were investigated in a national U.S. cohort (Wenten

et al., 20 09, 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 O 3 AQCD, examined the effects of ambient air p ollution
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, esp ecially 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

O 3 on acute effects, communities were divided into high and low exposure groups by median levels

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1 (46.9 ppb O 3 ). The epistatic effect of CAT and MPO variants w as evident in communities exhibiting



### 2 high ambient O 3 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 O 3 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 O 3 - related hospital discharges and emergency department (ED) visits for



### 9 asthma and respiratory disease that were available in the 2006 O 3 AQCD mainly looked at the da ily

time metric

. New studies evaluate d long

- term O 3 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 O 3 exposure.


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



### 14 declining warm

- season O 3 concentrations and hospitalization for asthma in children in California’s
South Coast Air Basin who ranged in age from birth to 19 y ea rs . Quarterly av erage concentrations



### 16 from 195 spatial grids, 10 × 10 km, were used

. Ozone was the only pollutant associated with



### 17 increased hospital admissions over the st udy 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 O 3 

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



### 22 2000. O zone was associated with reduced risk in the single

- pollutant model, but this relation did not



### 23 persist in multi -pollutant models.



### 24 I n a cross- sectional s tudy, Meng et al. (2010, 594252 ) examined associations between air
pollution and asthma morbid ity 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 O 3 based on resid ential 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 O 3 and PM did not differ substantially from single

- pollutant estimates

, indicating that pollutant multi

- collinearit y is not a problem in these analyses

. The authors report ed



### 31 increased asthma

- related ED visits or hospitalizations for O 3 (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 val ues. Associations for symptoms for adults (ages 18 +) were observed (OR 1.40 ; [ 95 % CI : 1.02

, 

### 34 1.91} per 10 ppb ).

A ssociations 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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35 respondent was a ssigned 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 mon ths

. T his cross

- sectional study reports an OR of 3.34 ( 95% CI : 1.01, 11.09 )
for poorly controlled asthma when comparing those 65 y ea rs of age and older above the 90th
percentile (28.7 ppb ) level to those below that l evel

. Multi -pollutant analysis produced similar

results.

E vidence associating long

- term O 3 exposure to first asthma hospital admission in a

concentration

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

196680 ) . This study investigated the association between chronic exposure to O 3 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 y ea rs 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 O 3 exposure: (1) mean concentration during the follow

- up period

(41.06 ppb) ; (2) mean concentration during the O 3 seaso n (50.62 ppb) ; and (3) proportion of fol low

up days with O 3 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 interes t. The effects of co -pollutant s 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 O 3 and childhood asthma
hospital ad missions was observed indicating that children exposed to high O 3 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 y ea rs) , 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 O 3 levels for all chronic exposure indicators (ORs, 1.16

- 1.68). When
estimating the O 3 effect using the excee dance proportion, a n increase was observed ( OR 1.68; [ 95%

CI : 1.64, 1.73]) in hospital admissions with an IQR (2.51%) increase in O 3 . 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 O 3 ( HR from the C ox 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 O 3 exposure to first asthma hospital admission in a concentration

- response relationship.

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### Figure 7-3

. O zone - 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 d uring 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: lo w [ 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 evide nce that long

- term exposure to ambient O 3 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

- ” O 3 communities

. O zone concentrations from the least to most polluted communities



### 7 (mean annual average of 8-h avg O 3 ) ranged from 30 to 65 ppb, whereas the ranges observed for the



### 8 other pollutants had four- to eightfol d differences in concentrations

. In a more recent CHS st udy,



### 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 associa ted 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 NO 2 , PM 10 , PM 2.5 ,

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39 elemental carbon, organic carbon, and O 3 . 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 FEV 1 , - 92.9 to

- 126.7 mL for FVC and

- 193.9 to

- 277.9 mL /s for MMEF for all
pollutants except O 3 , for which some positive associations were reported: 25.9 mL for FEV 1 ; 0.1 mL

for FVC, and 166.5 mL /s for MMEF. O zone 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 FEV 1 , -17.2 mL for FVC, and

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

As d iscussed in the 2006 O 3 AQ CD, a study of freshman students at the Universi ty of
California, Berkeley reported that lifetime exposure to O 3 was associated with decreased measures of
small airways ( <2 mm) function (FEF 75 and FEF 25

- 75 ) (Tager et al., 2005, 087538 ) . There was an

interaction with the FEF 25

- 75 / FVC ratio, a measure of intrinsic airway size. Subjects with a large ratio
were less likely to have decreases in FEF 75 and FEF 25

- 75 for a given estimated lifetime exposure to

O 3

. Kinney and Lippmann (2000, 011913 ) examined 72 nonsmoking adults (mean age 20 y ea rs )

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

O 3 exposures. Ilhorst et al. (2004, 055608 ) examined 2,153 children with a median age of 7.6 y ea rs and reported summer pulmonary function results which indicated that significantly lower FVC and

FEV 1 increases were associated with higher O 3 exposures in the summer, but not in the winter. Semi

annual mean O 3 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 O 3 levels for FVC and FEV 1 , 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 t o 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 y ea rs of life, the first 6 y ea rs 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 PM 10 , NO 2 , and CO during key neonatal and early life developmental periods. The
authors did not find a negative effect of exposure to O 3 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 PM 10 during the entire pregnancy and seco nd trimest er , and higher PM 10 during the first 2 years of life. Lower exposure to O 3 during the entire pregnancy or seco nd 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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39 contribute independently to different dimensions of school

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

- function measures (FVC, FEV 1 , PEF, FEF 25

- 75 ), prenatal exposures were more
influential on pulmonary function than early

- lifetime metrics, while, in contrast, the ratio of

measures (FEV 1 /FVC and FEF 25

- 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 a ny of the outcomes. This suggests that the timing of the O 3 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 O 3 and lung

function.

I n a cross

- sectional study of adults, Qian et al. (2005, 093283 ) e xamined the association of
long

- term exposure to O 3 and PM 10 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 O 3 exposure from daily data was determined at the

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

function.

T o determine the extent to which lo ng

- 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). FEV 1 was not associated with O 3 concentrations. In contrast

to the results for PM 10 , NO 2 , and SO 2; combin in g the results of all the survey years showed that a

5- ppb difference in O 3 was counter- intuitively associated with a high er FEV 1 by 22 mL

. In a prospective cohort study consisting of school

- age , non

- asthmatic children in Mexico City
(n = 3,170) who were 8 y ea rs of age at the beginning of the study, Rojas

- Martinez et al. (2007,

091064 ) evaluated the association between long

- term exposure to O 3 , PM 10 and NO 2 and lung

functi on growth every 6 mon ths 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 O 3 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 O 3 concentration of 11.3 ppb was associated with an annual deficit in FEV 1 of

12 mL in girls and 4 mL in boys. S ingle

- pollutant models showed an association between ambient
pollutants ( O 3 , PM 10 and NO 2 ) 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 child ren using passive O 3 samplers. O zone


### 2 concentrations were correlated (p < 0.05) with the measurements obtained from the fixed

- site air



### 3 monitoring stations.



### 4 In the 2006 O 3 AQ CD

, few studies had investigated the effect of chronic O 3 exposure on pulmonary fun ction. The strongest evidence was for seasonal effects of extended O 3 exposures on



### 6 lung function in children, i.e., reduced lung function growth being associated with higher ambient O 3 

### 7 levels. Longer

- term studies

, investigating the association of chronic O 3 exposure on yearly lung



### 8 function suc h as the CHS

, were inconclusive. Thus for new studies for pulmonary function, in one



### 9 study where O 3 and other pollutant levels were higher (90 ppb at high end of the range) than those in
the CHS, a relationship between O 3 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 th e other.



### 13 Thus there is little new evidence to build upon the very limited studies from the 2006 O 3 AQCD.



## 7.2.3.1. Evidence from Toxicological Studies

14 As r eviewed in the 1996 and 2006 O 3 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 expos ure 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, O 3 - inhalation studies have b een 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 O 3 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 O 3. (Carey et al., 2007, 195752 ; Chang MM -J ; Wu et al., 1998, 011983 ;


### 26 Chen et al., 2003, 035576 ; Duan e t 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 a l., 1985, 040278 ; Harkema et al., 1987, 041496 ; Harkema et a l., 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 O 3 AQCDs, the initial observations in adult non

- human primates



### 36 have been expanded in a series of experiments using infant rhesus monk eys repeatedly exposed to



### 37 0.5 ppm O 3 starting at 1 month of age (Plopper et al., 2007, 596412 )

. Many of the observations

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39 found in adult monke ys 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 monke ys to O 3 ( each cycle:
5 day s of 0.5 ppm O 3 at 8 h/day

, followed by 9 day s 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 inh aled histam ine (Schelegle et al., 2003, 053778 ), although neither
measurement was statistically different from filtered air contr ol 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 (O 3 + antigen) produced statistically significant and greater than additive changes in both functional

me asurements. This non

- human primate evidence

, of an O 3 - induced change in airway
responsiveness

, supports th e biologic plausibility of long_ term exposure to O 3 contributing to the

adverse effects of asthma in children

. To understand which conducting airways and inflammatory

mechanisms are involved in O 3 - induced airway hyperresponsiveness in the infant rhesus monkey, a

follow

- up study examined airway responsiveness ex vivo in lung sl ices (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 O 3 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
hyperrespo nsiveness in the large bronchi, whereas O 3 + 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 br onchioles.
The functional changes in the conducting airways of infant rhesus monkeys exposed to either

O 3 alone or O 3 + antigen were accompanied by a number of cellular and morpho lo gical changes,
including a significant fourfold increase in eosinophils, ( a cell type important in allergic asthma ) , in
the bronchoalveolar lavage of infant monkeys exposed to O 3 alone. Thus, these studies demonstrate
both functional and cellular changes in the lung of infant monkeys after cyclic exposure to 0.5 ppm

O 3 . This con centration, while higher than those used in controlled human exposure studies, provides
relevant information to understanding the adverse effects of ambient O 3 exposure on the respiratory

tract of humans. No concentration

- response data, however, are availa ble 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 O 3 . During normal
respiratory tract development, conducting airways increase in diameter and length in the infant
rhesus monkey. Exposure to O 3 alone (5 days of 0.5 ppm O 3 at 8 h/day

, followed by 9 days of
filtered air exposures for 11 cycles), however, markedly affected the growth pattern of distal

cond ucting airways (Fanucchi et al., 2006, 09649 1 ) . Whereas the first alveolar outpocketing
occurred at airway gene ration 13 or 14 in filtered air

- control infant monkeys, the most proximal

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39 alveolarized airways occurred at an average of 10 airway generations in O 3 - exposed monkeys.
Similarly, the diamet er and length of the terminal and respiratory bronchioles were significantly

decreased in O 3 - exposed monkeys. Importantly, the O 3 - induced structural pathway changes persisted
after recovery in filtered air for 6 mon ths after cessation of the O 3 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 inf ant rhesus monkeys exposed to O 3 + house dust mite antigen

, but not to O 3 alone (Schelegle et

al., 2003, 053778 ) . To s tudy the underlying mechanisms of this airway hyperresponsiveness, these
investigators evaluated the effect of exposure to O 3 alone and in combination with (+) antigen on

non

- specific airway responsiveness to methacholine at different airway generations. A fter exposure

to filtered air, O 3 , antigen, or O 3 + antigen, the bronchi and respiratory bronchioles of 6

- mo n th -old
monkeys were challenged ex vivo with methacholine. Exposure to O 3 alone had no significant effect
on airway responsiveness to methacholine i n either airway, whereas O 3 + 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 O 3 alone or O 3 + 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 O 3 alone had li ttle effect on mast cell numbers
and the response of animals exposed to O 3 + antigen was not different from that of animals exposed
to antigen alone ; thus suggesting that mast cells played little role in the interaction between O 3 and
antigen in this model of allergic asthma (Van Winkle et al., 2010, 670301 ) . Increases in CD4+ and
CD8+ lymphocytes were observed at 6 mon ths of age in the blood and bronchoalveolar lavage fluid
of infant rhesus monkeys exposed to O 3 + 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 O 3 + antigen. Although O 3 alone had no effect on CD25+ cells, it did alter the anatomic
distribution of CD25+ cells within the airways . Ozone had only a small e ffect 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 O 3 on allergic asthma.

In additio n 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 O 3 AQCD, exposure of infant rhesus monkeys altered the normal development of neural innervation in
the epithelium of the conducting airways (Larson e t al., 2004, 057062 ) . Whereas, a significant
reduction in airway innervation occurred after exposure to O 3 alone, a significantly greater reduction

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1 occurred in monkeys exposed to O 3 + 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 O 3 alone - and



### 4 O 3 + antigen- exposed infant monkeys after a 6

- mo n th recovery protocol (Kajekar et al., 2007,

567661 ) . Thus, in addition to structural, immune, and inflammatory effects, exposure to O 3 produces



### 6 alterations in airway innervation which may contribute to O 3 - induced exacerbation of asthma.



### 7 While the infant rhesus monkey studies examined the effect of long

- term O 3 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 O 3 . I n mice, age

- related differences in

O 3 - induced inflammation and the immediate

- early gene response were observed. Johnston and



### 11 colleagues (2006, 097439 ) demonstrated that the lung damage produced by O 3 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 a nd 2.5 ppm O 3 for 4 hour s) 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 O 3 (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 thro ughout the postnatal stage



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

- rel ated differences
in the pulmonary response of mice to O 3 . 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 O 3 for



### 23 5 hour s, this strain

- dependency suggests that genetic host factors play a role in age

- related



### 24 differences in response to O 3 . Thus, these rodent studies suggest that the response to O 3 in the
neonatal pe riod 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 O 3 exposure is



### 28 capable of damaging the distal airways an d 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 O 3 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 Ox idative Stres s

34 The 2006 O 3 AQCD state d that the extensive human clinical and animal toxicological
evidence, together with the limited epidemiologic evidence available, suggests a causal role for O 3 in



### 36 inflammatory responses in the airways. Though the major ity of recent studies focus on short

- term

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39 exposures, several epidemiologic and toxicology studies of long

- term exposure add to observations

of O 3 - induced inflammation and injury.

I nflammatory 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 day s after relocation to a rural location with significantly lower pollutant levels

(Renzetti et al., 2009, 199834 ) . The authors observe d a fourf old 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

, includ ing PM 10 , NO 2 , and O 3 , though
pollutant- specific resul ts were not presented. These results are consistent wi th 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 infl uence may be
rapidly reversible. E xhaled 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 day s between examinations for eNO needs to be evaluated for appropriateness, the results

suggest that inflammatory responses are reduced when O 3 levels are decreased

. C hest radiographs (CXR) of 249 children in Mexico City who were chronically exposed to O 3
and PM 2.5 were analyzed by Calderón

- Garcidueñas et al. (2006, 091253 ) . They report ed an association between chronic exposures to O 3 and other pollutants and a signi ficant increase in
abnormal CXR’s and lung CTs suggestive of a bronchiolar, per i bronchiolar, 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 cau tion because it is difficult to attribute effects to air
pollution e xposure.

I n a cross

- sectional study, Wood et al. (2009, 597085 ) examined the associat ion 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

- reso lution computed tomograph y 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 O 3 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 O 3 exposure was associated with worse gas transfer and more severe emphysema, albeit
accounting for only a small proporti on of the lung function variability. This suggests that a gene

specific group demonstrates a long

- term O 3 exposure effect.

The similarities of non

- human primates to humans make them attractive models in which to
study the effects of O 3 on the respiratory t ract. The nasal mucous membranes, which protect the more
distal regions of the respiratory tract, are susceptible to injury from O 3 . Carey et al. (2007, 195752 )
conducted a study of O 3 ex posure 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 O 3 , or

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1 episodically for several biweekly cycles alternating 5 days of 0.5 ppm O 3 with 9 days of filte red air



### 2 (0 ppm O 3 ), designed to mimic human exposure (70 days total). All monkeys acutely exposed to O 3 

### 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 expo sed continuously to 0.3 ppm O 3 in another study (Harkema et al., 1987, 040816 ) . A



### 8 90

- day exposure of rats to 0.8 ppm O 3 (8 h/day) elicited significantly elevated pro -inflammatory



### 9 eicosanoids PGE 2 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 associat ion of air pollutants with childhood respiratory allergies was examined i n 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 examine d the associations between the
reporting of respiratory allergy or hay fever and summer exposure to O 3 , controlling for



### 16 demographic an d geographic factors. Increased respiratory allergy/hay fever was associated with



### 17 increased O 3 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 poll utants, 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 ) report ed no relationship



### 22 between O 3 levels and rhinitis symptoms and hay fev er. 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 cross



### 24 sectional study in Taiwan. I n a large cross

- sectional study in France , Penard

- Morand et al. (2005,

087951 ) report ed a positive relationship between lifetime allergic rhinitis and O 3 exposure in a two



### 26 pollutant model with NO 2

. These studies relate d positive outcomes of allergic response and O 3 

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



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


### 29 after relocation from a highly polluted urban area in Italy to a rural location with significantly lower
pollutant level s (Renzetti et al., 2009, 199834 )

. 

### 31 T otal 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 a l., 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 h ome address. O zone concentrations were positively related to total IgE levels and an increase of 5 ppb of O 3 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 O 3 with NO 2 were decreased by 25% while NO 2 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 O 3 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 O 3 exposure and allergic outcomes.



## 7.2.6. Ho s t Defe ns e

8 Short- term exposures to O 3 cause decreases in host defenses against infectious lung disease in



### 9 animal models. However, acute O 3 - induced suppression of alveolar phagocytosis and immune
functions observed in animals appear s to be transient and attenuated with continuous or repeated



### 11 exposures. Chronic exposures (weeks, mont hs) 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 ) , n o detrimental effects were seen with a 120 -day exposure to 0.5 ppm O 3 on acute lung injury from influenza virus administered immediately before O 3 exposure started. But



### 16 there were O 3 - enhanced posti n fluenzal 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 O 3 and mortality. T he 2006 O 3 AQCD concluded that an insufficient amount of evidence
existed “to suggest a causal relationship between chronic O 3 exposure and increased risk for



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

- pulmonary mortality were



### 22 con sidered 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 O 3 elevated the risk of death from



### 26 respiratory causes and this effect was robust to the inclusion of PM 2.5 . The association between



### 27 increased O 3 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 severa l 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 O 3 AQCD detected no associations between



### 31 long

- term O 3 exposures and asthma

- related symptoms, asthma prevalence

, or allergy to com mon


### 32 aeroallergens among children after controlling for covariates. Little evidence was available to relat e

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### 1 long

- term exposure to current ambient O 3 concentrations to deficits in the growth rate of lung



### 2 function in children. Additionally, limited evidence w as available evaluating the relationship



### 3 between long

- term O 3 levels and pulmonary inflammation and other endpoints. From toxicological



### 4 studies, it appeared that O 3 - induced inflammation tapered off during long

- term exposures, but that



### 5 hyperplastic and fibr otic 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 O 3 

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



### 11 See Table 7- 1 for O 3 concentrations associated with selec ted studies. The positive results from



### 12 various designs and locations support an association between long

- term O 3 concentrations and



### 13 respiratory morbidity

. 

### 14 New studies examine d the relationship between long

- term O 3 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 O 3 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 th ey are methodologically rigorous epide miology studies. The studies were conducted in 12



### 20 California communities. The stratified analysis for the two independent fo u rth

- grade cohorts of the



### 21 study population recruited in 1993 and 1996 yielded consistent resul ts and provides replication in



### 22 independent groups of children

. Also, no meaningful interactions were observed between other air



### 23 pollutants such as PM 10 and genes.



### T

able 7-1. Summary of selected key new studies examining annual ozone exposure and respirator y health effects

Study; Health Effect; Location Mean Annual O 3 Concentration (ppb) O 3 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 a l. (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 O 3 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 O 3 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 a sthmatic children; CHS A bove 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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36

37 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 O 3 exposure and health effects in asthmatics. A long

- term O 3 exposure study relates bronchi tic

symptoms to TNF

- 308 genotype asthmatic children with ambient O 3 exposure in the CHS (Lee et al.,

2009, 199915 ) . A study relating asthma severity to long

- term O 3 exposure in five French cities
provides additional support to the notion that effects on as thma are related to long

- term O 3 exposure (Rage et al., 2009, 196720 )

. For the respiratory health of the general U.S. population, risk of
respiratory

- related sch ool absences was elevated for children with the CAT and MPO variant genes related to communities with high ambient O 3 levels (Wenten et al., 2009, 597084 )

. Chronic O 3 exposure was related to first childhood asthma hospita l admissions in a positive

conc entration

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

- sectional st udy in San Joaquin Valley in California reports similar
findings (Meng et al., 2010, 594252 ) . Another study relates asthma hospital admissions to quarterly

average O 3 in the South Coast Air Basin of California (Moore et al., 2008, 196685 ).
Information from toxicological studies indicates that long term exposure to O 3 during gestation
or development can result in irreversible morphological chan ges 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 O 3 (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 O 3 - induced change in airway responsiveness supports the biologic
plausibility of long term exposure to O 3 contributing to the adverse effects of asthma in children.

Results from e pidemiologic studies examining long

- term O 3 exposure and pulmonary function

effects are inconclusive with some new studies relat ing effects at higher exposure levels . T he results

from the CHS still remain as the definitive line of evidence . Other cross

- sec tional studies provide

mixed results.

The 2006 O 3 AQCD states that the extensive human clinical and animal toxicological
evidence, together with the limited epidemiologic evidence available, suggests a causal role for O 3 in
inflammatory responses in the a irways. Though the majority of recent studies focus on short

- term
exposures, several epidemiologic and toxicology studies of long

- term exposure add to observations

of O 3 - induced inflammation and injury. Toxicological studies in rodents and non

- human primat es

indicate that chronic O 3 exposure causes structural changes in the respiratory tract, and simulated
seasonal expos ure studies suggest that such exposures might have cumulative impacts. The strongest
epidemiologic evidence for a relationship between long

- term O 3 exposure and respiratory morbidity
is provided by new studies that demonstrate associations between long

- term measures of O 3

exposure and new

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

While there are current ly 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 O 3 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 - O 3 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 O 3 communities (mean = 38.4 ppb ) and from 46.5 to 64.9 ppb in high O 3 

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



### 8 long

- term O 3 exposure and respiratory morbidity effects such as first asthma hospitalization and



### 9 respiratory symptoms in asthmatics. Studies considering other pollutants provide data s u ggesting



### 10 that the effects related to O 3 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 da ta base provides a compelling case to support the



### 13 hypothesis that a relationship exists between long

- term exposure to ambient O 3 and me asures of



### 14 respiratory morbidity. The 2006 O 3 AQCD concluded the evidence was suggestive but inconclusive



### 15 at that time. Th e 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 c ausal



### 17 relationship between long

- term exposure to O 3 and respiratory morbidity.



# 7.3. Card iovas 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 O 3 and its effects on cardiovascular morbidity were not considered in



### 19 the 2006 O 3 AQCD (U.S. EPA, 2006, 088089 ) . However, recent studies have assessed the chronic



### 20 effects of O 3 exposure on cardiovascular morbidity (Chen et al., 2007, 145956 ; Ch uang et al., 2011,



### 21 670846 ; Forbes et al., 2009, 190351 )

. The association between O 3 exposure and markers of lipid



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



### 23 students, aged 18

- 22 y ea rs , 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 O 3 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 O 3 

### 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 O 3 

### 32 exposure was related to 8

- iso

- PGF [ β = 0.025 (pg/ mL )/8

- h ppb O 3 , p = 0.0007]. Fo r the 17

- ppb

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1 cumulative lifetime O 3 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 O 3 exposure and FRAP [ β =

- 2.21 (pg/ mL )/8

- h ppb O 3 , p = 0.45]. The



### 4 authors note that O 3 was highly correlated with PM 10

- 2.5 and NO 2 in this study population; however,

their inclusion in the O 3 models did not substantially modify the magnitude of the associations with



### 6 O 3 . Because the lifetime expo sure results were supported by shorter

- term exposure results from



### 7 analyses considering O 3 concentrations up to 30 days prior to sampling, the authors conclude that



### 8 persistent exposure to O 3 can lead to sustained oxidative stress and increased lipid peroxid ation.



### 9 However, because there was not much overlap in lifetime O 3 exposure estimates between LA and SF,
it is possible that the risk estimates involving the lifetime O 3 exposures could be confounded by



### 11 unmeasured factors related to other differences betwee n the two cities.



### 12 Forbes et al. (2009, 190351 ) u sed 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

f or 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 O 3 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 O 3 concentrations were 26.7 ppb, 25.4 ppb, and 28 ppb for 1994, 1998, and 2003,
respectively. Year speci fic adjusted effect estimates were created and combined in a meta -analysis.



### 21 No evidence of association was observed for O 3 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 O 3 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 O 3 exposure on cardiovascular morbidity risk.
A study was performed in Taiwan to exami ne the association between long

- term O 3 

### 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 O 3 concentration during the study period was 22.95 ppb



### 29 (SD 6.76 ppb). Positive associat ions were observed between O 3 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 mm Hg (95% CI : 18.14, 22.97) per 8.95 ppb increase in O 3 , respectiv ely).



### 32 Increased O 3 concentrations were also associated with increased levels of total cholesterol, fasting



### 33 glucose, hemoglobin A1c, and neutrophils. No associations were observed between O 3 

### 34 concentrations and triglyceride and IL

- 6 levels. The observed asso ciations 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 O 3 in



### 37 animal models. In addition to the short

- term effects des cribed in Section 6.3.1, a recent study found

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36 that A poE -/- mice (6

- 14 weeks old) exposed to O 3 (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 O 3 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 O 3 (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, endothel ial 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, O 3 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

O 3 exposure and is considered pro

- atherogenic. Both LOX

- 1 mRNA and protein were increased in

mouse aorta after O 3 exposure. This study provi des a possible pathway and further support to the

observed O 3 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. O zone 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 -infla mmatory mediators and decreasing production

of anti

- inflammatory proteins (Perepu et al., 2010, 385020 ) . Both long

- and short

- term O 3 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, O 3 induced oxid ative stress by decreasing SOD enzyme activity
and increasing malondialdehyde levels. O zone 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

O 3 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 O 3 and mortality. T he 2006 O 3 AQCD concluded that an insufficient amount of evidence



### 3 existed “to s uggest a causal relationship between chronic O 3 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 oppo sed to combined in the Pope



### 8 et al. (2002, 024689 ) work . A 10

- ppb increment in exposure to O 3 elevated the risk o f death from the



### 9 cardiopulmonary, cardiovascular, and ischemic heart disease. Inclusion of PM 2.5 as a co -pollutant
attenuated the association with exposure to O 3 for all of the cardiovascular endpoints to become null.



## 7.3.3. Summary and Causal De te rmina tio n

11 Prev ious AQCDs did not address the cardiovascular effects of long

- term O 3 exposure due to



### 12 limited data availability. The evidence remains limited; however the emerging data is supportive of a



### 13 role for O 3 in chronic cardiovascular diseases. Few epidemiologic st udies have investigated



### 14 cardiovascular morbidity after long

- term O 3 exposure , and the majority only assessed cardiovascular
disease related biomarkers. A study on O 3 and cardiovascular mortality reported no association after



### 16 adjustment for PM 2.5 levels. Fu rther epidemiologic studies on cardiovascular morbidity and mortality



### 17 after long

- term exposure have not been published.



### 18 Toxicological evidence on long

- term O 3 exposure is also limited but three strong toxicological



### 19 studies have been published since the previous AQCD. These studies provide evidence for O 3 enhanced atherosclerosis and I/R injury, corresponding with development of a systemic oxidative,



### 21 proinflammatory environment. Although questions exist for how O 3 inhalation causes systemic



### 22 effects, a recen t study proposes a mechanism for development of vascular pathology that involves



### 23 activation of LOX

- 1 by O 3 oxidized lipids and proteins. This activation may also be responsible for



### 24 O 3 induced changes in genes involved in proteolysis, thrombosis, and vasoco nstriction. Taking into
consideration the positive toxicological studies reported, the generally limited body of evidence is



### 26 s uggestive of a causal relationship between relevant long

- term exposures to O 3 and



### 27 cardiovascular effects.



# 7.4. Reproductive and Develo pmental 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 num ber of studies, v arious 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 ai r pollution research

, while congenital malforma tions are less studied

. 

### 3 Infants and fetal development processes may be particularly vulnerable to O 3 exposure, and



### 4 although the physical mechanisms are not fully understood, several hypotheses have been propo sed
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 O 3 AQCD examined impacts on

birth

- related endpoints, including intrauterine, perinatal, postneonatal, and infant deaths; premature



### 11 births; intrauterine gr owth 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 O 3 , no associations were found between O 3 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 ch allenges 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 suc h 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 (Ka vlock et al.,



### 23 1980, 094043 ) found that exposure to O 3 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 O 3 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

correspo nd 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 ri sk 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. hav e also reported associations



### 2 between ambient air pollution and sperm quality for individual air pollutants, including O 3 and



### 3 PM 2.5 . 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 O 3 in healthy sperm



### 6 donors. This effect persis ted under a joint additive model for O 3 , CO, NO 2 and PM 10 . The authors did



### 7 not detect a reduction in sperm count. Hansen et al. (2010, 594438 ) investigated the effect of



### 8 exposure to O 3 and PM 2.5 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 con centration and



### 13 count with O 3 , 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 O 3 - induced oxidative stress, inflammatory reac tions, 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 test icular morphology have been demonstrated in



### 18 toxicological studies with histological evidence of O 3 - 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 spermat id cells, and focal epithelial desquamation with denudation to the



### 22 basement membrane). The exposure protocol used five month old adult rats exposed to O 3 as adults



### 23 (0.5 ppm, 5 h/d ay for 50 days). This degeneration could be rescued by vitamin E administrati on,


### 24 indicating an antioxidant effect. Vitamin C administration had no effect at low doses of ascorbic acid

and exacerbated the O 3 - 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 O 3 exposure that was rescued with certain



### 28 antioxidant supplementation.



### 29 Overall, there is limited epidemiologic evidence for an association with O 3 concentrat ion 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 expo sure 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

o utside 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 co ncentrations of criteria pollutants at addresses of women undergoing their first IVF cycle and at



### 3 their IVF labs from 2000 to 2007 in the north eastern U.S. (Legro et al., 2010, 597377 ) . Increasing O 3 

### 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 pp b 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 NO 2 in a co

- pollutant model, however, O 3 was


### 8 no longer significantly associated with IVF failure. The results of this study suggest that exposure to



### 9 O 3 during ovulation was beneficial (perhaps due to early conditioning to O 3 ), whereas later exposure

to O 3 (e.g., during gestation) was detrimental, and reduced the likelihood of a live bir th.



### 11 In toxicological studies, reproductive success in rats appears unaffected by O 3 exposure

. 

### 12 Ozone administration (continuous 0.4, 0.8 or 1.2 ppm O 3 ) 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 O 3 ). Initially, dam body weight (0.8 and



### 17 1.2 ppm), water consumption (0.4, 0.8 and 1.2 ppm O 3 ) and food consumption (0.4, 0.8 and 1.2 ppm)



### 18 during pregnancy were decreased with O 3 exposure but these deficits dissipated a week or two after



### 19 the initial exposure (Bignami et al., 1994, 076063 ) . The anorexigenic effect of O 3 exposure on the
pregnant dam appears to dissipate with time; the dams seem to a dapt to the O 3 exposure. In males,



### 21 data exist showing morphological evidence of altered spermatogenesis in O 3 exposed animals



### 22 (Jedlinska- Krakowska et al. (2006, 195640 ). Some evidence suggests that O 3 may affect



### 23 reproductive success when combined with other chemicals. Kavlock et al. (1979, 039228 ) showed



### 24 that O 3 acted synergistically with sodium salicylate to increase the rate of pup resorptions after
midgestational exposure (1.0 ppm O 3 , GD9

- 12). At low doses of O 3 exposure, toxicological studies



### 26 show r eproductive effects to include a transient anorexigenic effect of O 3 on gestational weight gain,



### 27 and a synergistic effect of O 3 on salicylate

- induced pup resorptions; other fecundity, pregnancy and



### 28 gestation related outcomes appear unaffected by O 3 exposur e. Collectively, there is very little



### 29 epidemiologic evidence for the effect of O 3 on reproductive success, and the reproductive success in
rats appears to be unaffected in toxicological studies of O 3 exposure.



## 7.4.3. Birth Weigh t

31 With birth weight routinely coll ected 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 O 3 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 O 3 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) O 3 concentrations. This effect



### 9 was stronger for concentrations averaged over the second and third trimesters. PM 10 , NO 2 and CO



### 10 concentrations averaged over the entire pregnancy were not statistically significa ntly associated with



### 11 birth weight, though CO concentrations in the first trimester and PM 10 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 O 3 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 O 3 , estimates for the next 5 lowest deciles were approximately

- 40 g to

- 50 g, with



### 16 no clear trend and with 95% confidence bounds t hat included zero. The highest four deciles of O 3 

### 17 exposure showed an approximately linear decrease in birth weight, and all four 95% CIs excluded



### 18 zero , and r anged from mean decreases of 74 grams to dec reases of 148 grams.

Source: Salam et al. (2005, 087885 )



### Figure 7- 4. Birthw eight deficit by deci l e 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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39 Several additional studies conducted in the U.S. and Canada also investigated the association

between ambient O 3 concentrations a nd birth weight and found little evidence for an association.

Morello- Frosch et al. (2010, 670076 ) estimated ambient O 3 concentratio ns 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 O 3 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 O 3
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 O 3 concentrations to create exposure variables based on average maternal exposure for each trimester.

O zone was not found to be related to birth weight in single

- pollutant models, though the O 3 effect
during the third trimester was borderline significant in a co -pollutant model with PM 10 . 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 monito ring 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 ass ociations between

elevated concentrations of CO and PM 10 both early and late in pregnancy and risk of term LBW.
After adjusting for CO and/or PM 10 the authors did not observe associations between O 3 and term
LBW in any of their models. Brauer et al. (2008, 156292 ) evaluated the impacts of air pollution (CO,

NO 2 , NO, O 3 , SO 2 , PM 2.5 , PM 10 ) on birth weight for the period 1999

- 2002 using spatiotemporal

resi dential exposure metrics by month of pregnancy in Vancouver, BC. Quantitative results were not
presented for the association between O 3 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 amon g a cohort of term singleton births

from 1988

- 2000. Though there was some indication of an association with SO 2 and PM 10 , there were

no effects for O 3

. Similarly, studies conducted in Australia, Latin America, and Asia report limited evidence for

an associ ation between ambient O 3 and measures of birth weight. In Sydney, Australia, Mannes et al.

(2005, 087895 ) found that O 3 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 sta tistically
significantly associated with a reduction in birth weight in Brisbane, Australia. In Sao Paulo, Brazil,

Gouveia et al. (2004, 055613 ) found that O 3 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 O 3 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 O 3 effect during the first trimester of pregnancy, but they found that during the third



### 3 trimester of pregnancy O 3 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 O 3 ( 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 ) north ern 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 from monitor 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 Near est 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 O 3 ^Odds ratios of LBW; Highest quartile of exposure compared to lowest quartile of exposure +Relative risk of L BW per 10 ppb change in O 3 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 stro ng evidence for an effect of ambient O 3 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 St udy. The toxicological literature on the effect of O 3 on birth weight is sparse. In some studies, the
reporting of birth weight may be avoided bec ause birth weight can be confounded by decreased litter



### 6 size resulting from an increased rate of pup resorp tio n (aborted pups) in O 3 exposed dams. In one



### 7 toxicological study by Haro and Paz (1993, 044194 )

, no diff erences in litter size were observed and



### 8 decreased birth weight in pups f rom dams who were exposed to 1ppm O 3 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 neo natal 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 exposur e 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 eff ect. 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 determina nt . The studies reviewed have dealt with this issue in



### 24 dif ferent 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 week s ) 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 week s is a 32

- week fetus 4 week s prior to



### 29 birth, while an infant born a t term (40 w ee ks ) is a 36

- week fetus 4 w ee ks 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 c oncentrations during the days immediately preceding birth. An advantage of these time



### 33 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 -contro l 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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39 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 O 3 exposure and PTB. These include: decreased in utero oxygen supply leading to a reduction of oxygen carrying
capacity; changes in blo od 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 refe rence (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, altho ugh 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 O 3 . Mos t 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, NO 2 , SO 2 , O 3 , PM 10 , PM 2.5 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 O 3

. 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, NO 2 , O 3 , PM 10 ) exposure during pregnancy on the occurrence of PTB in a cohort of 97,518 neonates born in

south ern 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 O 3 over any of the
pregnancy periods in single o r multi

- pollutant models. Wilhelm and Ritz (2005, 088668 ) extended
previous analyses of PTB (Ritz and Yu, 1999, 0 86976 ; 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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39 heterogeneity inadequately captured by ambient monitors. The authors analyzed the association

between O 3 exposure during varying periods of pregnancy and PTB, finding a positive association

between O 3 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 O 3 ) and the first month of pregnancy (results for first trimester exposure were
similar, but slightly s maller, quantitative results not presented) in models containing all pollutants.

No association was observed between O 3 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 effec t
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 PM 2.5 , no consistent patterns of increase in the odds of preterm birth for

O 3 or NO 2 were observed.
One study conducted in Canada evaluated the impacts of air pollution (including CO, NO 2 , NO, O 3 , SO 2 , PM 2.5 , and PM 10 ) on preterm births (1999

- 2002) using spatiotemporal residential

exposu re 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 PM 2.5 for PTB. The O 3 associations were largely protective, and no quantitative results
were presented for O 3 . Additionally, Lee et al. (2008, 195720 ) used time

- series techniques to
investigate the short

- term associations of O 3 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 betw een ambient O 3 and PTB. Hansen et al. (2006, 089818 ) reported that exposure to O 3 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 O 3 during the first trimester. The effect estimate was diminished and lost statistical significance when

PM 10 was included in the model (OR 1.23, [ 95% CI : 0.97, 1.59 ] per 10 ppb increase). Maternal

exposure to O 3 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 O 3 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 O 3 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 O 3 

### 2 correspond ed to 9.47 % (95% CI : 0.70, 18.7%) increase in PTBs. Increases in PTB were also



### 3 observed for PM 10 , SO 2 , and NO 2 . 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 wee k 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 O 3 . This



### 8 phenomenon was most likely due to the anorexigenic ef fect of relatively high O 3 concentrations



### 9 (Bignami et al., 1994, 076063 ).



### Table 7

- 3. Br ief summary of epidemiologic studies of PTB

Study

Darrow et al. (2009, 195818 ) Location Sample Size

Atlanta, GA (n=476,489) Mean O 3 (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 e t 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 pre gnancy: 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)

Jal aludin 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 bir th: 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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### 5

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30 #Relative risk of P TB per 10 ppb change in O 3

. * Relative risk of PTB per 1 ppb change in O 3

. 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 report ed



### 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 O 3 and reports no



### 3 association with PTB. However when long

- term exposure to O 3 early in pregnancy is examined the



### 4 results are inconsistent. Studies conducted in the U.S., Canada, and England find no association with

O 3 and PTB, while studies conducted in Australia and China report an O 3 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 10 th 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 environment al 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 record ed



### 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 g rowth 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 unkno wn. 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,

cer tain 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 placenta tion 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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39 intrauterine growth via vascular dysfunction in the placenta and damaged DNA. Also, pro

inflammatory cytokines can limit trophoblast invasion during the early stages of pregnancy,
restricting fetal growth. Poor placental vascularity is caused partly by dysregulation of gene
expressio n 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 <10 th 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 overest imate 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 10 th
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 ar e 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 O 3 concentrations on IUGR in a
population of infants born in California from 1975

- 1987 as part of the Children’s Health Study. Th e authors reported that maternal O 3 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 O 3 exposure during the third trimester increased the risk of I UGR by 11% (95% CI: 0, 20%). Brauer et

al. (2008, 156292 ) evaluated the impacts of air pollution (CO, NO 2 , NO, O 3 , SO 2 , PM 2.5 , PM 10 ) on

SGA (1999

- 2002 ) using spatiotemporal residential exposure metrics by month of pregnancy in

Vancouver, BC. The O 3 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 O 3 . Liu et al. (2007, 090429 ) examined the association between IUGR among singleton

term live births and SO 2 , NO 2 , CO, O 3 , and PM 2.5 in 3 Canadian cities for the period 1985

- 2000. No
increase in the risk of IUGR in relation to exposure to O 3 averaged over each month and trimester of
pregnancy was noted.
Three studies conducted in Australia provide evidence for a n association between ambient O 3 and fetal growth restriction. Hansen et al. (2007, 090703 ) examined SGA among singleton, full

- term

births in Brisban e, Australia in relation to ambient air pollution (bsp, PM 10 , NO 2 , O 3 ) 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 sign ificantly 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 O 3 during month 4 (OR 1.11 [95% CI: 1.00, 1.24 ] per 10 ppb increase). In a subsequent study,

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34 Hansen et al. (2008, 190273 ) examined the possible associations between fetal ultrasonic

mea surements and ambient air pollution (PM 10 , O 3 , NO 2 , SO 2 ) 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 measurem ents, 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

associ ated with decreases in all of the fetal growth measurements, and a 1.78 mm reduction in
abdomen circumference per 10 ppb increase in O 3 concentration, though this effect did not persist in

co

- pollutant models. The change in ultrasound measurements associat ed with O 3 during days 31 -60
of gestation indicated that increasing O 3 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 a l. (2005, 087895 ) estimated the effects of pollutant (PM 10 , PM 2.5 , NO 2 , CO and O 3 )
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 O 3 in the second trimester of pregnancy had

small but adverse effects on SG A (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 O 3 can affect postnatal
growth. A few studies reported that mice or rats exposed developmentally (gestationally ±

la ctationally) to O 3 had deficits in body weight gain in the postpartum period (Bignami et al., 1994,

076063 ; Haro and Paz, 1993, 044194 ; K avlock et al., 1980, 094043 )

. Table 7

- 4 provides a brief overview of the epidemiologic studies of fetal growth restriction. In
summary, the evidence is inco nsistent when examining exposure to O 3 and fetal growth restriction.
Similar to PTB, studies conducted in Australia have reported an effect of O 3 on fetal growth, whereas
studies conducted in other areas have not found such an effect. This may be due to th e 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 O 3 (ppb) Effect Estim ate (95% CI) ( Sample Size ) a ssessment

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 1 56292 ) (n=70,249) Inverse Distance T1: NR Weighting (IDW) T3: NR Entire pregnancy: NR (results Calgary, Edmonton, and presented in fig ure) 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, Bri sbane, 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 O 3

. * 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 O 3 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 va rious 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 develo pment . 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. O zone 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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39 Several studies hav e been conducted examining the relationship between O 3 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 O 3 ), p ulmonary artery and valve anomalies (OR 1.34 [ 95% CI: 0.96, 1.87 ] per 10 ppb O 3 ),

and conotruncal defects (OR 1.36 [ 95% CI: 0.91, 2.03 ] per 10 ppb O 3 ) in a dose

- response manner

with second

- mo nth O 3 exposure. These associations were not observed for orofaci al defects. The
average effect sizes and patterns of second

- mo nth O 3 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. Addition ally, 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 ( O 3 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 O 3 exposure and isolated ventri cular 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 pro vide strong evidence that air pollution increases the risk of cardiac defects or

oral clefts.
A recent study conducted in Atlanta, GA examined O 3 exposure during the third through

seventh week of pregnancy and reported no association with risk of cardiovas cular malformations
(mean long

- term average of 8 -h O 3 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 O 3 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 O 3 exposure were st able 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 exam ined
and cleft malformations, though there was a weak association between cases of cleft palate only and
increasing O 3 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 O 3 m onitor (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 O 3 . Kavlock et al. (1979, 039228 ) exposed pregnant rats to O 3 for precise periods



### 5 during organogenesis. No significant teratogenic effects were found in rats exposed 8 hr/day to



### 6 concentrations of O 3 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 O 3 (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 associatio ns between O 3 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 O 3 and card iovascular 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 O 3 (ppb) Exp osure A ssessment Exposure W indow

Ritz et al. (2002, 023227 ) Cardiac and Cleft Defects south ern California (n=3,549 cases; 10,649 controls) 24

- h avg: NR Nearest Monitor (with in 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=65 3 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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39 asthmatic children (ages 6

- 11) included in the Fresno Asthmatic Children’s Environment Study
(FACES). Monthly me ans 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 y ea rs of life, the first 6 y ea rs of life, and the entire l ifetime. The 8

- h avg O 3 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 PM 10 , NO 2 , and CO during key neonatal and early life developmental periods.

The authors di d not find a negative effect of exposure to O 3 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 PM 10 during the entire pregnancy and seco nd trimester and
higher PM 10 during the first 2 years of life. Lower exposure to O 3 during the entire pregnancy or

seco nd 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, FEV 1 , PEF, FEF 25

- 75 ), prenatal
exposures were more influential on pulmonary function than early

- lifetime metrics, while, in

contrast, the ratio of measures (FEV 1 /FVC and FEF 25

- 75 /FVC) were most in fluenced 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 ca se

- 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

calcul ated for the duration of pregnancy (~15 ppb; transformed from μg /m 3 ) and the first year of life

(~14 ppb; transformed from μg /m 3 ). In contrast to the Mortimer et al. studies (2008, 122163 ; 2008,

187280 ) , the effect estim ates for first

- year exposure were generally larger than for in utero
exposures. However, similar to the Mortimer et al. studies, the observed associations with O 3 were largely protective. Because of the relatively high correlation between in utero and firs t- year exposures for many pollutants, it was difficult to discern the relative importance of the individual
exposure periods.

Latzin et al. (2009, 195721 ) examined whe ther 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 O 3 concentrations averaged across the post

- natal period was ~44 ppb

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39 (transformed from μg/m 3 ). Consistent with the previous studies, no association was found for
prenatal exposure to O 3 and lung function.
The new toxicological literature since the 2006 O 3 AQCD (U.S. EPA, 2006, 088089 ) , covering
respiratory changes related to deve lopmental O 3 exposure, reports ultrastructural changes in
bronchiole development, alterations in placental and pup cytokines, and increased pup airway hyper

reactivity. 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 t o 1

- ppm O 3 (12 h/d ay , 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 O 3 had
cytoplasmic damage and swollen mitocho ndria. Bronchial epithelium at the canalicular phase in O 3 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, dela yed
maturation of tracheal epithelium following early neonatal O 3 exposure (1 ppm, 4

- 5 h/day for

fir st 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

O 3 exposed rat fetuses. Finally, in the saccular phase, mitochondrial degradation was present in the

non

- ciliated bronchiolar cells of rats exposed in utero to O 3 . In conclusion, O 3 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 differenti ally 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 O 3 - induced airway hyper

- reactivity in
juvenile offspring. Maternal PM exposure also significantly increased placental cytokines above

vehicle

- instilled controls. Pup postnatal O 3 exposure ( 1 ppm 3 h/d ay , every other day, thrice week ly

for 4 week s ) induced significantly increased cytokine levels (IL

- 1β, TNF

- α , KC, and IL

- 6) in whole
lung versus pos tnatal 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 O 3 starting at one

- mo nth of age have examined the effect of O 3 alone or in combination with an inhaled
allergen on morphology and lung function (Plopper et al., 2007, 596412 ) . Exposure to O 3 alone or allergen alone produced small but not statistically significant changes in baseline airway resistance
and airway responsiveness, but the combined exposure to both O 3 + 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 O 3 (Fanucchi et al., 2006

, 096491 ). A more detailed description of these

studies can be found in S ection 7.2.3 (Pulmonary Structure and Function).

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## 7.4.8. De velop me n tal Central Nervous Sys te m Effects

7.4.8.1. Laterality

1 Two reports of laterality changes in mice developmentally exposed to O 3 have been reported in



### 2 the literature. Mice developmentally exposed to 0.6 ppm O 3 (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 spe cific structural asymmetries of the hippocampal mossy fiber projections (Schöpke



### 6 et al., 1991, 684254 ). The 2006 AQCD evidence for the effect of O 3 on laterality or handedness



### 7 demonstrated that rats exposed to O 3 during fetal and neonatal life showed limited, gender

- specific



### 8 changes in handedness after exposure to the intermediate dose of O 3 (only seen in female mice



### 9 exposed to 0.6 ppm O 3 , and not in males at 0.6 ppm or in either sex of 0.3 or 0.9 ppm O 3 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 O 3 (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 s ynthesis; 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 O 3 (Boussouar et al., 2009, 596368 ) when compa red to control air exposed

animals who also were restrained. The O 3 - exposed offspring in this study were cross

- fostered to



### 21 control air exposed dams to avoid O 3 - dependent dam related neonatal effects on offspring outcomes



### 22 (i.e., dam behavioral or lactationa l contributions to pup outcomes) (Boussouar et al., 2009, 596368 )

. 

### 23 Developmental exposure to 0.3 or 0.6 ppm O 3 prior to mating pair formation through GD17



### 24 induced significant increased levels of BDNF in the striatum of adult (PND140) O 3 exposed
offspring as compared to control air exposed animals; these O 3 - 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 O 3 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 (Romero



### 31 Velazquez et al., 2002, 035575 ) , and altered monoamine neurotransmitter content with the



### 32 catecholamine system affected and the indoleamine system unaffected by O 3 (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 O 3 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.

O 3 administration (0.4, 0.8 or 1.2 ppm O 3 ) during the majority of pregnancy (PD7

- 17) of CD

- 1



### 6 mice did not affect p up 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 - O 3 exposed dams

. Testing for aggressive behavior in mice continuously exposed to O 3 (0.3 or 0.6 ppm from



### 11 30 day s 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 ver sus air exposed controls (Santucci et al., 2006, 596414 )

. Similar to this and as reported in



### 14 previous AQCDs, continuous exposure of adult animals to O 3 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 to tal 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 O 3.



## 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 O 3 exposure (1 ppm O 3 , 12h/d ay , 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 O 3 exposed



### 29 offspring spent longer time in the wakefulness state during the light pe riod, more time in SWS
during the period of darkness, and showed significant decrements in PS. Chronic O 3 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 aberratio ns in adult animals exposed to O 3 [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 O 3 - dependent sleep

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1 changes potentially exists with evidence from a publication on indomethacin pretreatment



### 2 attenuating O 3 - 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 s usceptible. Several age intervals have been explored: neonatal



### 7 (<1 month ); postneonatal (1 month to 1 y ea r) ; and an overall interval for infants that includes both



### 8 the neonatal and postneonatal periods (<1 y ea r). Within these various age categories, multip le 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. A s discussed below, a handful of studies have examined the effec t of



### 14 ambient air pollution on neonatal and postneonatal mortality

, with the former the least studied. These
studies varied somewhat with regard to the outcome s 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 pro ximate 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 determina nt . 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 NO 2 , but lesser for SO 2 and CO. These



### 31 associations exhibited a short lag time, less than 5 days. No significant assoc iation was detected



### 32 between O 3 and intrauterine mortality.

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## 7.4.9.2. Infa nt Mo rtality, Les s than 1 Y ear

1 Ritz et al. (2006, 089819 ) linked birth and death certificates for infants who died between



### 2 1989 and 2000 to evaluate the i nfluence of outdoor air pollution on infant death in the South Coast



### 3 Air Basin of California. The authors examined exposure periods 2 w ee ks, 1 mon th , 2 mon ths , and



### 4 6 mon ths before each case subject's death and reported no association between ambient levels of O 3
and infant mortality. Similarly, Diaz et al. (2004, 089894 ) analyzed the effects of extreme



### 6 temperatures and air pollutants on daily mortality in children le ss than 1 y ea r of age in Madrid,



### 7 Spain, from 1986 to 1997 and observed no statistically significant association between mortality and



### 8 O 3 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 O 3 

### 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 NO 2 and O 3 3-5 days before death, but not as consistently as with PM. A 10

- ppb increase in 24

- h avg

O 3 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 two



### 17 pollutant model with PM 2.5

. One

- hour max concentrations of O 3 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 M onth

19 Three studies have evaluated ambient O 3 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 w ee ks, 1 mon th , 2 mon ths,



### 23 and 6 mon ths before each case subject's death and reported no association between ambient levels of



### 24 O 3 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 b etween



### 27 O 3 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

O 3 and neonatal mortality.



## 7.4.9.4. Pos tne o na ta l Morta lity, 1 Mon th to 1 Y ear

31 A number of studies focused on the postneonatal period when examining the effects of O 3 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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38 Coast Air Basin of California. The authors examined exposure periods 2 w ee ks, 1 mon th , 2 mon ths

, and 6 mon ths before each case subject's death and reported no association between ambient levels of

O 3 and postneona tal mortality. Woodruff et al. (2008, 098386 ) evaluated the county

- level
relationship between cause

- specific postneonatal infant mortality and chronic early

- life (first

2 m on ths of life) exposure to air pollutants across the U.S. Similarly, they found no association

between O 3 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 t here was no relationship between O 3 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 O 3 concen trations 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 O 3 lev els with postneonatal infant mortality in

the Mexico City Metropolitan Area between 1997 and 2005. Generally, O 3 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 O 3 was significantly related to respiratory mortality
among those with low SES. In a separate analysis, the effect of PM 10 was evaluated with O 3 level
quartiles. PM 10 alone was related to a significant increase in all

- cause mortality . The magnitude of
this effect remained the same when only the days when O 3 was in the lowest quartile were included
in the analyses. However, when only the days when O 3 was in the highest quartile were included in
the analyses, the magnitude of the PM 10 e ffect increased dramatically (OR=1.06 [0.909, 1.241] for

PM 10 on days with O 3 in lowest quartile; OR=1.26 [1.08, 1.47] for PM 10 on days with O 3 in the
highest quartile. These results suggest that while O 3 alone may not have an effect on infant mortality,

i t may serve to potentiate the observed effect of PM 10 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 O 3 . 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 O 3 concentrations and infant mortality in the postneonatal period.

Ha e t al. (2003, 042552 ) conducted a daily time

- series study in Seoul, Korea to evaluate the effect of

short- term changes in ambient 8 -h O 3 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 case

crossover analysis to examine the relationship between air pollution exposure and postneonatal
mortality in Taipei, Taiwan for the period 1994

- 2000. The authors observed no association s between

ambient levels of O 3 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 betw een ambient O 3 concentrations and SIDS comes
from a study that evaluated the county

- level relationship between SIDS and chronic early

- life (first

2 mon ths 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 O 3 and deaths from
SIDS. There was a monotonic increase in odds of SIDS for each quartile of O 3 exposu re compared
with the lowest quartile (highest quartile OR = 1.51; [ 95% CI: 1.17, 1.96 ] ). In a multi -pollutant
model including PM 10 or PM 2.5 , CO and SO 2 , the OR for SIDS and O 3 was not substantially lower
than that found in the single

- pollutant model. When examined by season, the relationship between

SIDS deaths and O 3 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 O 3 and the OR for normal weight babies was 1.16 (95% CI: 1.01,

1.32) per 10

- ppb increase in O 3

. Conversely, two additional studies reported no association between ambient levels of O 3 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 w ee ks, 1 mon th , 2 mon ths , and

6 mon ths before each case subject's death and reported no association between ambient levels of O 3
and SIDS. Dales et al. (2004, 087342 ) used time

- series analyses to compa re 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 SO 2 , NO 2 , and CO, but not O 3 o r PM 2.5

.

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### Table 7

- 6. Brief summary of infant mortality studies

Study Location Mean O 3 (ppb) Exposure
A ssessment 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)

L 1: 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 ) south ern 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 C itywide 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-h a 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 O 3 * 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 repor ted

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

The 2006 O 3 AQCD concluded that the limited number of studies that investigated O 3
demonstrated no associations between O 3 and birth outcomes, with the possible exception of birth

defects. The current review included an expanded body of evidence on the associations between O 3 and reproductive and developmental effects. Recent epidemiologic and toxicological studies provide
evidence for an effect of prenatal exposure to O 3 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 O 3 on sperm quality, and prov ide limited evidence for decrements in sperm concentration, while
there is limited toxicological evidence for testicular degeneration associated with O 3

. While the collective evidence for many of the birth outcomes examined is generally

inconsistent (incl uding birth defects), there are several well

- designed, well

- conducted studies that

indicate an association between O 3 and adverse outcomes. For example, as part of the south ern California Children's Health Study, Salam et al. (2005, 087885 ) observed a concentration

- response relationship of decreasing birth weight with increasing O 3 co ncentrations 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 O 3 during days 31

- 60 of gestation indicated that increasing O 3 concentration
decreased an ultrasound measurement for women living within 2 km of the monitoring site.
There is no evidence that prenatal or early life O 3 concentrations are associated with infant
mortality. Collectively, there is limited though positive toxicological evidence for O 3 -induced
developmental effects, including effects on pul monary structure and function and central nervous
system effects. Limited epidemiologic evidence for an effect on prenatal O 3 exposure on respiratory
development provides coherence with the effects observed in toxicological studies. There is also

limited e pidemiologic evidence for an association with O 3 concentration and decreased sperm
concentration. A recent toxicological study provides limited evidence for a possible biological
mechanism (histopathology showing impaired spermatogenesis) for such an assoc iation.
Additionally, though the evidence for an association between O 3 concentrations and adverse birth
outcomes is generally inconsistent, there are several influential studies that indicate an association
with reduced birth weight and restricted fetal g rowth. Taking into consideration the positive evidence
for developmental and reproductive outcomes from tox i cological and epidemiological studies, and

the few influential birth outcome studies, the evidence is suggestive of a causal relationship between

lo ng - term exposures to O 3 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 O 3 exposure and neurobehavioral effec ts. Chen et al. (2009, 179945 ) utilized data from the NHANES
III cohort to study the relationship between O 3 levels (mean annual O 3 concentrati on 26.5 ppb) and
neurobehavioral effects among adults aged 20

- 59 y ea rs . The authors observed an association
between annual exposure to O 3 and tests measuring coding ability (symbol

- digit substitution test)

and attention/short

- term memory (serial

- digit lear ning test). Each 10

- ppb increase in annual O 3 levels corresponded to an aging

- related cognitive performance decline of 3.5 yr for coding ability

and 5.3 y ea rs for attention/short

- term memory. These associations persisted in both crude and
adjusted models. There was no association between O 3 levels and reaction time tests. The authors
conclude that overall, there is an association between long

- term O 3 exposure and reduced
performance on neurobehavioral tests.
In a subchronic study, rats were exposed to 0.25 ppm O 3 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 O 3 (4 h/day) for 30 or 6 0 days (Guevara

- Guzmán et al., 2009, 596385 )

. In the olfactory
bulb, lipid peroxidation was si gnificantly less in rats exposed to O 3 and tre ated daily with estradiol.
This protective effect of estradiol was also demonstrated for O 3 - induced decrements in a selective
olfactory recognition memory test and an ol factory- dependent reward test. Similarly, estradiol
protected against O 3 - induced changes in nigral cell morphology and loss of dopamine neurons in rats
exposed to O 3 fo r 30 days (Angoa

- Pérez et al., 2006, 596366 ). Thus, repeated exposure of rats to O 3 produces lipid peroxidation at multiple sites in the brain and this oxidative stress is accompanied by
gene expression changes and decrements in behavioral test s. Olfactory changes and loss of
substantia nigra neurons are associated with Parkinson's disease in humans. Inhibition of these

effects with estradiol treatm ent 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 O 3 occurred in ut ero

. Several neurotransmitters were assessed in male offspring of dams
exposed to 1 ppm O 3 during the entire pr egnancy (Gonzalez

- Pina et al., 2008, 475317 )

. The data
showed that catecholamine neurotransmitters were affected to a greater degree than indole

- amine

neurot ransmitters in the cerebellum. Adverse CNS changes, including behavioral, c ellular, and

biochemical effects, have also been observed after in utero exposure to 0.5 ppm O 3 for 12 h/day from

GD 5 to GD 20 (Boussouar et al., 2009, 596368 )

. Tyrosine hydroxylase labe ling in the nucleus tractus

soli tarius was increas ed after in utero exposure to O 3 whereas Fos pr otein 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 O 3 offspring

. Although the effec t of O 3 exposure


### 3 concentration was not studied in these two in utero studies, it h as 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 O 3 . I ncreased defensive/submissive behavior and reduc ed



### 6 social inves tigation were observed in both the 0.3

- and 0.6

- ppm O 3 groups. Changes in gene



### 7 expression of brain

- derived neurotrophic factor (BDNF, increased in striatum) and nerve growth



### 8 factor (NGF, decreased in hippocampus) accompa nied these behavioral changes. Thus, these three



### 9 studies demonstrate that CNS effects can occur as a result of in utero exposure to O 3 , 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 O 3

. 

### 14



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

15 The 2006 O 3 AQCD (U.S. EPA, 2006, 088089 ) included toxicological evidence that acute



### 16 exposures to O 3 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 neurobehavio ral effects was not



### 19 available. R ecent r esearch in the area of O 3 - induced neurotoxicity has included several long

- term



### 20 exposure studies. Notably, the first epidemiologic study to examine the relationship between O 3 

### 21 exposure and neurobehavioral effects obser ved an association between annual O 3 levels and an



### 22 aging- related cognitive performance decline in tests measuring coding ability and attention/short



### 23 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 O 3 . Additionally, neurobehavioral changes are evident in animals



### 26 whose only exposure to O 3 occurred in utero . Collectively, the lim ited epidemiologic and



### 27 toxicological evidence is coherent and suggestive of a causal relationship between O 3 exposure



### 28 and adverse CNS effects.



# 7.6. Carcinogenic and Genotoxic Potential of O zone

7.6.1. Introd uc tion

29 The radiomimetic and clastogenic qualities of O 3 , combined with its ability to stimulate



### 30 proliferation of cells in the respiratory tract, have suggested that O 3 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 O 3 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 O 3 as a pulmonary carcinogen” (U.S. EPA,



### 3 2006, 088089 )

. 

### 4 Multiple epidemiologic studies reported in the 2006 O 3 AQCD examined the direct association

between O 3 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 O 3 and lung cancer mortality

. The Adventist Health



### 8 Study of Smog (AHSMOG) also examined the association between O 3 and lung cancer mortality



### 9 (Abbey et al., 1999, 047559 )

. There was a positive a ssociation between O 3 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 O 3 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 O 3 concentrations



### 16 and lung cancer mor tality due to low variability in O 3 levels across the cities (Dockery et al., 1993,



### 17 044457 )

. An ecologic study performed in Sa o Paulo City, Brazil examined the correlations between



### 18 O 3 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 O 3 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 l ung adenoma 1 acceleration in mice with daily



### 23 exposure to 1 ppm over 15 mon ths (Stokinger, 1962, 015101 ) . Later work demonstr ated 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 O 3 (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 t o 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, 04250 9 ) . A study of carcinogenicity of O 3 by the



### 33 National Toxicology Program (NTP, 1994, 011143 ) reported increased i ncidences of



### 34 alveolar/bronchiolar adenoma or carcinoma (combined) in female B6C3F 1 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 NO TE: 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. Fro m page 8

- 33 of the 1970 O 3 AQCD: “No true lung cancers have been reported, however, from experimental exposures
to either O 3 alone or any other combination or ingredient of photochemical oxidants.”

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1 evidence of carcinogenic activity of O 3 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 O 3 

### 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 O 3 was reported in a study of male



### 6 Sprague- Dawley rats exposed for 6 mon ths (n = 50) (Monchaux et al., 1996, 086197 ) . These two



### 7 studies also examined co

- carcinogenicity of O 3 with NNK 1 (Boorman et al., 1994, 038195 ) or a


### 8 relat ively high dose of radon (Monchaux et al., 1996, 086197 ) , finding no enhancement of NNK



### 9 related tumors and a slight non

- significant increase in tumo r 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 O 3 could potentially delay or inhibit tumor development (Witschi et al., 1993,



### 13 043206 ) . Thus there is no concrete evidence that O 3 can act as a co

- carcinogen.



### 14 Immune surveillance is an imp ortant defense against cancer, and it should be noted that n atural

killer (NK) cells, which destroy tumor cells in the lung, appear to be inhibited by higher doses of O 3 

### 16 and either unaffected or stimulated at lower doses (S ection 6.2.5.4, Infection and Ada ptive



### 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 O 3 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 In cid 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 O 3 and cancer mortality since the last

O 3 AQCD

. Krewski et al. (2009, 191193 ) c onducted an extended follow

- up of the cohort



### 26 (1982

- 2000)

. Mean O 3 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 O 3 (HR 1.00 [ 95% CI: 0.96 -1.04]



### 29 per 10 ppb O 3 )

. Additionally, no association was observed when O 3 was restricted to the summer

months

. There was also no association present in a sub

- analysis of the cohort examinin g the



### 31 relationship between O 3 and lung cancer mortality in the Los Angeles area.


### 32 Since the 2006 O 3 AQCD, two toxicological studies have examined potential carcinogenicity



### 33 of O 3 (Kim and C ho, 2009, 200775 ; Kim and Cho, 2009, 200773 ) . Looking acr oss both studies,



### 34 which used the same mouse strain as the National Toxicology Program study described above,

0.5 ppm O 3 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 O 3 for 16 week s. 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 carcinom a after 32 week s 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 O 3 exposure was predicted from the



### 7 radiomimetic properties of O 3 . The decomposition of O 3 in water produces OH and HO 2 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 O 3 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 O 3 , 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 hour s . Other in vivo
exposure studies found increased DNA strand breaks in respiratory cells from guinea pigs (Ferng et



### 16 al., 1997, 083578 ) and mi ce (Bornholdt et al., 2002, 03667 7 ) but only with exposure to high er doses



### 17 of O 3 (1 ppm for 72 hour s and 1 or 2 ppm for 90 min utes , 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 m ice exposed to 0.2 ppm during gestation (blepharoph imosis 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, O 3 was found to be mutagen ic in Salmonella

, 

### 22 with and w ithout S9 metabolic activation. No new toxicological studies of DNA damage have



### 23 become available since the 2006 O 3 AQCD.



### 24 A number of epidemiologic studies looked at the association between O 3 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 O 3 

### 28 exposures. Huen et al. (2006, 089035 ) examined cytogenetic damage among Africa n American



### 29 children and their mothers in Oakland, CA. Increased O 3 (mean monthly 8 -h O 3 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 bu ccal 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 O 3 over the summer. However, O 3 levels over 2, 7, 10, 14, or 30 days



### 37 ( concentrations not given ) before collection of buccal cells did not correlate with cytogeneti c

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Additionally, the authors exposed a subset of the students (n=15) to 200 ppb O 3 for 4 hour s 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 hour s after exposure.
A study performed in Mexico re cruited 55 male workers working indoors (n=27) or outdoors

(n=28) in Mexico City or Puebla, Mexico in order to study the relationship between O 3 and DNA
damage (detected from peripheral blood samples using the Comet assay) (Tovalin et al. , 2006,

091322 ) . The median estimated daily O 3 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 O 3 levels and DNA
damage was observed. However, when examining the relationship by city and workplace, only DNA
damage in outdoor workers i n Mexico City remained correlated with O 3 levels.
Three studies examining the relationship between O 3 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 Prospect ive 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 y ea rs prior to enrollment) were enrolled in the study

. Th e researchers

examined DNA adduct levels (DNA bonded to cancer

- causing chemicals) and their relationship with

O 3 concentrations (concentrations not given) (Peluso M Hainaut et al., 2005, 089057 )

. A positive

association was seen between DNA adduct levels and O 3 concentra tions from 1990

- 1994 but not O 3
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 O 3 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 (deter mined by Comet assay using blood lymphocytes) in
association with varying periods of O 3 exposure

. The researchers observed that longer periods of
high O 3 exposure (concentrations not given) were more strongly correlated with oxidative DNA
damage than short er exposures ( i.e.

, the rho [p-value] was 0.26 [0.03] for 0

- 10 day s 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 days

. When adjusted for age, gender, smoking history, ≤ 25
traffic pollution exposure, period of blood draw, and area of residence, the association between O 3 levels and oxidative DNA damage was positive for O 3 levels 0

- 60 day s, 0

- 75 day s, and 0

- 90 day s 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 bein g

non

- smokers or light smokers (Giovannelli et al., 2006, 1 99894 )

. The estimated O 3 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

. O zone concentrations were correlated with DNA strand breaks



### 2 (measured from blood lymphocytes) over longer exposure periods (p

- value: 0.002 at 30 day s,


### 3 p- value: 0.04 at 7 days; p- value: 0.17 at 3 day s)

. This association was robust to control for



### 4 temperature, solar radiation, gender, and age

. No association was seen between O 3 concentratio ns and measures of oxidative DNA damage at 3, 7, or 30 day s.



## 7.6.4. S u mma ry an d Cau s al De te rmina tio n

6 The 2006 O 3 AQCD (U.S. EPA, 2006, 088089 ) reported that evidence did not support ambient



### 7 O 3 as a pulmonary carcinogen. Since the 2006 O 3 AQCD, very few epidemiologic and toxicological



### 8 studies have been published that examine O 3 as a c arcinogen, but collectively, study results indicate



### 9 that O 3 may contribute to DNA damage. Overall, the evidence is inadequate to determine if a
causal relationship exists between ambient O 3 exposures and cancer.



# 7.7. Mortality

11 A limited number of epidemiologic studies have assessed the relationship between long

- term



### 12 exposure to O 3 and mortality in adults

. T he 2006 O 3 AQCD concluded that an insufficient amount of



### 13 evidence existed “to suggest a causal relationship between chronic O 3 exposure and increased risk



### 14 fo r mortality in humans” (U.S. EPA, 2006, 088089 ) . In addition to the infan t 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 O 3 , and a reanalysis of the



### 17 ACS cohort that specifically points to a relationship between long

- term O 3 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 O 3 , (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 O 3 concentrations in six cities: Topeka, KS; St. Louis,



### 23 MO; Portage, WI; Harriman, TN; Steubenville, OH; and Watertown, MA. Mean O 3 concentrations



### 24 from 1977 to 1985 ranged from 19.7 ppb in Watertown to 28.0 ppb in Portage. Long

- term mean O 3 concentrations were not found to be associated with mortality in the six cities. However, the authors



### 26 noted that “The small differences in O 3 levels among the (six) cities limited the power of the study to



### 27 detect associations between mortality and O 3 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 P revention Study II). All

- cause, cardiopulmonary , lung cancer



### 32 and other mortality risk estimates for long

- term O 3 exposure are shown in Figure 7-5

. While no



### 33 consistently significant positive associat ions were observed between O 3 and mortality, the mortality



### 34 risk estimates were larger when analyses considered more accurate exposure metrics, rising when the

March 20 11 7

- 59 DRAFT – DO NOT CITE OR QUOTE

### 1 entire period was considered ( versus just at the start of the study) and becoming marginally



### 2 sign ificant 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 PM 2.5 were observed for both lung cancer and cardio

- pulmonary mortality.

Years of Data Number of Number of Participants 1- h Max O 3 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 O 3 , 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 O 3 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 O 3 exposure (mean O 3 concentration 26.2 ppb ) and fatal coronary heart disease

. Thus, no association



### 15 of chronic O 3 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 O 3 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 sm aller subcohorts based on exposure and mortality follow

- up periods. Four separate exposure

March 20 11 7

- 60 DRAFT – DO NOT CITE OR QUOTE

### 1

2

3

4

5

6

7

8

9

10

11

12

13

14

15

16

17

18

19

20

21

22

23

24

25

26

27

28

29

30

31

32

33

34

35

36

37

38

39 periods were associated with three mortality follow

- up periods

. For concurrent exposure periods,
peak O 3 was positively associated with all

- cause mortality, with a 9 .4% (95% CI: 0.4, 18.4) excess
risk per mean 95th percentile O 3 less estimated background level (not stated). “Peak” refers, in this
case, to the 95th percentile of the hour ly 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 O 3 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 o f O 3 ,” reporting a significant O 3 effect even with traffic density included in
the model: RR=1.080 per 40 ppb peak O 3 (95% CI : 1.019, 1.146)

. In Lipfert (2006, 088756 ) , which
considers only the EPA Speciation Trends Network (STN) sites, O 3 drops to non

- significant predictor
of total mortality for this cohort

. However, the authors acknowledge that: “Peak O 3 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 rela tively
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 NO 2 , O 3 and SO 2 , but PM w as not addressed

. The results

indicat ed that long

- term exposure to SO 2 was associated with cardio

- respiratory mortality, but the
fact that this study considered only one city, and that the range of O 3 exposure across that city

(23.7-35.6 ppb ) was low and slight in variation in comparison to t he 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 rec ent 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 O 3 concentration ranged from 33.3 to 104.0 ppb ) . A 10

- ppb
increment in exposure to O 3 elevated the risk of death from the cardiopulmonary, cardiovascular,

ischemic heart disease, and r espiratory causes. Inclusion of the concentration of PM 2.5 measured in

1999

- 2000 as a co

- pollutant attenuated the association with exposure to O 3 for all end points except
death from respiratory causes, for which a significant association persisted ( T able 7.7). The

association between increased O 3 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

March 20 11 7

- 61 DRAFT – DO NOT CITE OR QUOTE

### 1 considered individually. Subgroup analyses showed that temperature and region of country, but not



### 2 sex, age at enrollment, body

- mass index, education, or PM 2.5 concentration, modified the effects of



### 3 O 3 on th e risk of death from respiratory causes (i.e., risks were higher at higher temperature, and in



### 4 the Southeast, Southwest, and Upper Midwest). O zone threshold analyses indicated that the



### 5 threshold model was not a better fit to the data (p > 0.05) than a line ar representation of the overall



### 6 O 3 - mo rtality association. Overall, this new analysis indicates that long

- term exposure to PM 2.5 

### 7 increases risk of cardiac death, while long

- term exposure to O 3 is specifically associated with an



### 8 increased risk of respirator y 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 O 3 (96 MSAs) O 3 (86 MSAs) O 3 +PM 2.5 (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 D isease 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 T he 2006 O 3 AQCD concluded that an insufficient amount of evidence existed “to suggest a



### 10 causal relationship between chronic O 3 exposure and increased risk for mortality in humans”



### 11 (U.S. EPA, 2006, 088089 ) . Two additional studies have been conduc ted since the last review, an



### 12 ecologic study that finds no association between mortality and O 3 (Wang et al., 2009, 199990 ) , and a


### 13 re

- analysis of the ACS cohort that specifically points to a relationship between long

- term O 3 

### 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 short



### 17 and long

- term exposure to O 3 on respiratory effects. Additionally, the evidence for short

- and long



### 18 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 O 3 exposures


### 20 and all

- cause mortality.

March 20 11 7

- 62 DRAFT – DO NOT CITE OR QUOTE

# References

A list of all epidemiologic references considered for inclusion in this chapter can be found at http://hero.epa.gov/index.cfm?action=search.do&submit=Search&group_id=403

A li st of all toxicological references considered for inclusion in this chapter can be found at http://hero.epa.gov/index.cfm?action=search.do&submit=Search&group_id=401

A list of all controlled human exposure references considered for inclusion in this chapter can be found at http://hero.epa.gov/index.cfm?action=search.do&submit=Se arch&group_id=477

Abbey, D. E.; Nishino, N.; McDonnell, W. F.; Burchette, R. J.; Knutsen, S. F.; Beeson, W. L.; Yang, J. X. (1999). Long

term inhalable particles and other air pollutants related to mortality in nonsmokers. Am J Respir Cri t Care Med,

159: 373 -382. 047559

Akinbami, L. J.; Lynch, C. D.; Parker, J. D.; Woodruff TJ (2010). The association between childhood asthma prevalence

and mon itored air pollutants in metropolitan areas, United States, 2001

- 2004. Environ Res, 110: 294 -301. http://dx.doi.org/10.1016/j.envres.2010.01.001 378580

Angoa- Pérez, M.; Jiang, H.; Rodríguez, A. I.; Lemini, C.; Levine, R. A.; Rivas

- Arancibia, S. (2006). Estrogen counteracts

ozone- induced oxidative stress and nigral neuronal death. Neuroreport, 17: 629 -633. http://www.ncbi.nlm.nih.gov/entrez/query.fcgi?cmd=Retri eve&db=PubMed&dopt=Citation&list_uids=16603924.

596366

Arito, H.; Uchiyama, I.; Yokoyama, E. (1992). Acute effects of ozone on EEG activity, sleep

- wakefulnes s and heart rate in rats. Ind Health, 30: 23 -34. 042759

Auten, R. L.; Potts, E. N.; Mason, S. N.; Fischer, B.; Huang, Y.; Foster, W. M. (2009). Maternal expos ure to particulate matter increases postnatal ozone

- induced airway hyperreactivity in juvenile mice. Am J Respir Crit Care Med, 180:

1218

- 1226. http://dx.doi.org/10.1164/rccm.200901 -0116OC 200760

Beeson, W. L.; Abbey, D. E.; Knutsen, S. F. (1998). Long

- term concentrations of ambient air pollutants and incident lung cancer in California adults: results from the AHSMOG study. Environ Health Perspect, 106: 813

- 82 3. 048890

Berhane, K.; Gauderman, W. J.; Stram, D. O.; Thomas, D. C. (2004). Statistical issues in studies of the long

- term effects of air pollution: The southern California children’s health study. Stat Sci, 19: 414 -449. http://dx.doi.org/10.1214/088342304000000413 626732

Bignami, G.; Musi, B.; Dell'Omo, G.; Lavi ola, G.; Alleva, E. (1994). Limited effects of ozone exposure during pregnancy on physical and neurobehavioral development of CD

- 1 mice. Toxicol Appl Pharmacol, 129: 264 -271. http://dx.doi.org/10.1006/taap.1994.1251 076063

Bobak, M. (2000). Outdoor air pollution, low birth weight, and prematurity. Environ Health Perspect, 108: 173

- 176.

011448

Boorman, G. A.; Hailey, R.; Grumbein, S.; Chou, B. J.; Herbert, R. A.; Goehl, T.; Mellick, P. W.; Roycroft, J. H.; Haseman,

J. K.; Sills, R. (1994). Toxicology and carcinogenesis studies of ozone and ozone 4

- (N

- nitrosomethylami no)-1- (3

­ pyridyl) -1- butanone in Fischer

- 344/N rats. Toxicol Pathol, 22: 545

- 554. 038195

Bornholdt, J.; Dybdahl, M.; Vogel, U.; Hansen, M.; Loft, S.; Wallin, H . (2002). Inhalation of ozone induces DNA strand breaks and inflammation in mice. DNA Repair, 520: 63 -71. 036677

Boussouar, A.; Araneda, S.; Hamelin, C.; Soul age, C.; Kitahama, K.; Dalmaz, Y. (2009). Prenatal ozone exposure abolishes stress activation of Fos and tyrosine hydroxylase in the nucleus tractus solitarius of adult rat. Neurosci Lett, 452:

75-78. http://www.ncbi.nlm.nih.gov/entrez/query.fcgi?cmd=Retri eve&db=PubMed&dopt=Citation&list_uids=19444957.

596368

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).

March 20 11 7

- 63 DRAFT – DO NOT CITE OR QUOTEBracken, M. B.; Triche, E. W.; Belanger, K.; Saftlas, A.; Beckett, W. S.; Leaderer, B. P. (2003). Asth ma symptoms, severity, and drug therapy: a prospective study of effects on 2205 pregnancies. Obstet Gynecol, 102: 739

- 753. 156288

Brauer, M.; Lencar, C.; Tam buric, L.; Koehoorn, M.; Demers, P.; Karr, C. (2008). A cohort study of traffic

- related air pollution impacts on birth outcomes. Environ Health Perspect, 116: 680 -686. 156292

Breton, C. V.; Salam, M. T.; Vora, H.; Gauderman, W. J.; Gilliland, F. D. (2011). Genetic variation in the glutathione synthesis pathway, air pollution, and children's lung function growth. Am J Respir Crit Care Med, 183: 243 -248. http:// dx.doi.org/10.1164/rccm.201006- 0849OC 687660

Calderón

- Garcidueñas, L.; Mora

- Tiscareno, A.; Fordham, L. A.; Chung, C. J.; Valencia

- Salazar, G.; Flores

- Gomez, S.; Solt,

A. C.; Gomez

- del Campo, A.; Jardon

- Torres, R.; Henriquez

- Roldan, C.; Hazucha, M. J.; Reed, W. (2006). Lung radiology and pulmonary function of children chronically exposed to air pollution. Environ Health Perspect, 114:

1432 -1437. 091253

Carbajal

- Arroyo, L.; Miranda

- Soberanis, V.; Medina

- Ramón, M.; Rojas

- Bracho, L.; Tzintzun, G.; Solís

- Gutiérrez, P.;

Méndez

- Ramírez, I.; Hurtado

- Díaz, M.; Schwartz, J.; Romieu, I. (In Press). Effect of PM10 and O3 on infant mortality among residents in the Mexico City Metropolitan Area: A case

- crossover analysis, 1997

– 2005. J Epidemiol Community Health. http://dx.doi.org/10.1136/jech.2009.101212 667773

Carey, S. A.; Minard, K. R.; Trease, L. L.; Wagner, J. G.; Garcia, G. J.; Ballinger, C. A.; Kimbell, J. S.; Plopper, C. G.; Corley, R. A.; Postlethwait, E. M.; Harkema, J. R.; Einstein, D. R. (2007). Three

- dimensional mapping of ozone

induced injury in the nasal airways of monkeys using magnetic resonance imaging and morphometric techniques. Toxicol Pathol, 35: 27

- 40. http://dx.doi.org/10.1080/01926230601072343 195752

Chang MM

- J; Wu, R.; Plopper, C. G.; Hyde, D. M. (1998). IL

- 8 is one of the major chemokines produced by monkey airway epithelium after ozone

- induced injury. Am J Phys iol, 275: L524

- L532. 011983

Charpin, D.; Pascal, L.; Birnbaum, J.; Armengaud, A.; Sambuc, R.; Lanteaume, A.; Vervloet, D. (1999). Gaseous air pollution and atopy. Clin Exp Allergy, 29: 1474 -1480. 015152

Chen, C.

- Y.; Bonham, A. C.; Plopper, C. G.; Joad, J. P. (2003). Plasticity in respiratory motor control: selected contribution: neuroplasticity in nucleus tractus solitarius neurons following episodic ozone exposure in infant primates. J Appl Physiol, 94: 819 -827. 035576

Chen, C.; Arjomandi, M.; Balmes, J.; Tager, I.; Holland N (2007). Effects of Chronic and Acute Ozone Exposure on Lipid Peroxidation and Antioxidant Capacity in Healthy Young Adults. Environ Health Perspect, 115: 1732 -1737. http://dx.doi.org/10.1289/ehp.10294 145956

Chen, C.; Arjomandi, M.; Qin, H.; Balmes, J.; Tager, I.; Holland, N. (2006). Cytogenetic damage in buccal epithelia and peripheral lymphocytes of young healthy individuals exposed to ozone. Mutagenesis, 21: 131

- 137. http://dx.doi.org/10.1093/mutage/gel007 196504

Chen, J.

- C.; Schwartz, J. (2009). Neurob ehavioral effects of ambient air pollution on cognitive performance in US adults. Neurotoxicology, 30: 231

- 239. http://dx.doi.org/10.1016/j.neuro.2008.12.011 179945

Chen, L. H.; Knutsen, S. F.; Shavlik, D.; Beeson, W. L.; Petersen, F.; Ghamsary, M.; Abbey, D. (2005). The association between fatal coronary heart disease and ambient particulate air pollution: Are females at greater risk? Environ Health Perspect, 113: 1723 -1729. 087942

Chen, L.; Yang, W.; Jennison, B. L.; Goodrich, A.; Omaye, S. T. (2002). Air pollution and birth weight in northern Nevada,

1991 -1999. Inhal Toxicol, 14: 141 -157. 024945

Chuang, G. C.; Yang, Z.; Westbrook, D. G.; Pompilius, M.; Ballinger, C. A.; White, R. C.; Krzywanski, D. M.; Postlethwait,

E. M.; Ballinger, S. W. (2009). Pulmonary ozone exposure induces vascular dysfunction, mitochondrial damage, and atherogenesis. Am J Physiol Lung Cell Mol Physiol, 297: L209

- L216. http://dx.doi.org/10.1152/ajplung.00102.2009 197202

Chuang, K. J.; Yan, Y. H.; Chiu, S. Y.; Cheng, T. J. (2011). Long

- term air pollution exposure and risk factors for cardiovascular diseases among the elderly in Taiwan. Occup Environ Med, 68: 64 -68. http://dx.doi.org/10.1136/oem.2009.052704 670846

Clark, N. A.; Demers, P. A.; Karr, C. J.; Koehoorn, M.; Lencar, C.; Tamburic, L.; Brauer, M . (2010). Effect of early life exposure to air pollution on development of childhood asthma. Environ Health Perspect, 118: 284 -290. http://dx.doi.org/10.1289/ehp.0900916 594440

March 20 11 7

- 64 DRAFT – DO NOT CITE OR QUOTEClifton, V. L.; Giles, W. B.; Smith, R.; Bisits, A. T.; Hempenstall, P. A.; Kessell, C. G.; Gibson, P. G. (2001). Alterations of placental vascular function in asthmatic pregnancies. Am J Respir Crit Care Med, 164: 546 -554. 156360

Cole, M. P.; Freeman, B. A. (2009). Promotion of cardiovascular disease by exposure to the air pollutant ozone. Am J Physiol Lung Cell Mol Physiol, 297: L209

- L216

. 597507

Conti, D. V.; Cortessis, V.; Molitor, J.; Thomas, D. C. (2003). Bayesian modeling of complex metabolic pathways. Hum Hered, 56: 83

- 93. http://dx.doi .org/10.1159/000073736 626696

Dales, R.; Burnett, R. T.; Smith

- Doiron, M.; Stieb, D. M.; Brook, J. R. (2004). Air pollution and sudden infant death syndrome. Pediatrics, 113: 628 -631. 087342

Darrow, L. A.; Klein, M.; Flanders, W. D.; Waller, L. A.; Correa, A.; Marcus, M.; Mulholland, J. A.; Russell, A. G.; Tolbert

, P. E. (2009). Ambient air pollution and preterm birth: a time

- series analysis. Epidemiology, 20: 689

- 698. 195818

Dell'Omo, G.; Wolfer, D.; Alleva, E.; Lipp , H.

- P. (1995). Developmental exposure to ozone induces subtle changes in swimming navigation of adult mice. Toxicol Lett, 81: 91 -99. 080787

Diaz, J.; Linares , C.; Garcia

- Herrera, R.; Lopez, C.; Trigo, R. (2004). Impact of temperature and air pollution on the mortality of children in Madrid. J Occup Environ Med, 46: 768 -774. 089894

Dockery, D. W.; Pope CA 3rd; Xu, X.; Spengler, J. D.; Ware, J. H.; Fay, M. E.; Ferris, B. G. Jr; Speizer, F. E. (1993). An association between air pollution and mortality in six US cities. N Engl J Med, 329: 1753 -1759. 044457

Duan, X.; Buckpitt, A. R.; Pinkerton, K. E.; Ji, C.; Plopper, C. G. (1996). Ozone

- induced alterations in glutathione in lung subcompartments of rats and monkeys. Am J Res pir Cell Mol Biol, 14: 70 -75. 080791

Duan, X.; Buckpitt, A. R.; Plopper, C. G. (1993). Variation in antioxidant enzyme activities in anatomic subcompartments within rat and rhesus monkey lung. Toxicol Appl Pharmacol, 123: 73 -82. 086326

Dugandzic, R.; Dodds, L.; Stieb, D.; Smith

- Doiron, M. (2006). The association between low level exposures to ambient air pollution and term low birth weight: A retrospective cohort study. Environ Health, 5: 3. http://dx.doi.org/10.1186/1476- 069X -5-3 088681

Ercan, H.; Birben, E.; Dizdar, E. A.; Keskin, O.; Karaaslan, C.; Soyer, O. U.; Dut, R.; Sackesen, C.; Besler, T.; Kalayci, O. (2006). Oxidative stress and genetic and epidemiologic determinants of oxidant injury in childhood asthma. J

Al lergy Clin Immunol, 118: 1097 -1104. 595172

Evans, M. J.; Fanucchi, M. V.; Baker, G. L.; Van Winkle, L. S.; Pantle, L. M.; Nishio, S. J.; Schelegle, E. S.; Ge rshwhin, L. J.; Miller, L. A.; Hyde, D. M.; Sannes, P. L.; Plopper, C. G. (2003). Atypical development of the tracheal basement membrane zone of infant rhesus monkeys exposed to ozone and allergen. Am J Physiol, 285: L931

- L939. 048167

Evans, M. J.; Fanucchi, M. V.; Baker, G. L.; Van Winkle, L. S.; Pantle, L. M.; Nishio, S. J.; Schelegle, E. S.; Gershwin, L. J.; Miller, L. A.; Hyde, D. M.; Plopper, C. G. (2004). The remodelled tracheal basement membrane zone of infant rhesus monkeys after 6 months of recovery. Clin Exp Allergy, 34: 1131

- 1136. http://dx.doi.org/10.1111/j.1365

­ 2222.2004.02004.x CEA2004 596379

Fanucchi, M. V.; Plopper, C. G.; Evans, M. J.; Hyde, D. M.; Van Winkle, L. S.; Gershwin, L. J.; Schelegle, E. S. (2006). Cyclic exposure to ozone alters distal airway development in infant rhesus monkeys. Am J Phy siol Lung Cell Mol Physiol, 291: L644

- L650. http://dx.doi.org/10.1152/ajplung.00027.2006 096491

Fanucchi, M. V.; Wong, V. J.; Hinds, D.; Tarkington, B. K.; Van Winkle, L. S.; Evans, M. J.; Plopper, C. G. (2000). Repeated episodes of exposure to ozone alters postnatal development of distal conducting airways in infant rhesus monkeys [Abstract]. Am J Respir Crit Care Med, 161: A615. 012284

Feichtinger, W.; Papalambrou, K.; Poehl, M.; Krischker, U.; Neumann, K. (1997). Smoking and in vitro fertilization: A

meta

- analysis. J Assist Reprod Genet, 14: 596

- 599. http://dx.do i.org/10.1023/A:1022584802711 625251

Ferng, S.

- F.; Castro, C. E.; Afifi, A. A.; Bermudez, E.; Mustafa, M. G. (1997). Ozone

- induced DNA strand breaks in guine a pig tracheobronchial epithelial cells. J Toxicol Environ Health, 51: 353

- 367. 083578

Forbes, L. J. L.; Kapetanakis, V.; Rudnicka, A. R.; Cook, D. G.; Bush, T.; Stedman, J. R.; Whincup, P. H.; Strachan, D. P.; Anderson, H. R. (2009). Chronic exposure to outdoor air pollution and lung function in adults. Thorax, 64: 657

­ 663. 595425

March 20 11 7

- 65 DRAFT – DO NOT CITE OR QUOTEForbes, L. J.; Patel, M. D.; Rudnicka, A. R.; Cook, D. G.; Bush, T.; Stedman, J. R.; Whincup, P. H.; Strachan, D. P.; Anderson, R. H. (2009). Chronic exposure to outdoor air pollution and markers of systemic inflammation. Epidemiology, 2 0: 245

- 253. http://dx.doi.org/10.1097/EDE.0b013e318190ea3f 190351

Frischer, T.; Studnicka, M.; Gartner, C.; Tauber, E.; Horak, F.; Veiter, A.; Spengler, J.; Kuhr, J.; Urbanek, R. (1999). Lung function growth and ambient ozone: a three

- year population study in school children. Am J Respir Crit Care Med,

160: 390 -396. 001037

Frischer, T.; Studnicka, M.; Halmerbauer, G.; Horak, F.; Gartner, C.; Tauber, E.; Koller, D. Y. (2001). Ambient ozone exposure is associated with eosinophil activation in healthy children. Clin Exp Allergy, 31: 1213 -1219. 019683

Fujinaka, L. E.; Hyde, D. M.; Plopper, C. G.; Tyler, W. S.; Dungworth, D. L.; Lollini, L. O. (1985). Respiratory bronchiolitis following long

- term ozone exposure in bonnet monk eys: a morphometric study. Exp Lung Res, 8:

167-190. 040278

Gauderman, W. J. (2001). Sample size requirements for matched case

- control studies of gene -environ ment interaction. Stat Med, 21: 35

- 50. http://dx.doi.org/10.1002/sim.973 625862

Gauderman, W. J. (2002). Sample size requirements for association studies of gene- gene interaction. Am J Epidemiol, 155:

478-484. 626945

Gauderman, W. J.; Avol, E.; Gilliland, F.; Vora, H.; Thomas, D.; Berhane, K.; McConnell, R.; Kuen zli, N.; Lurmann, F.; Rappaport, E.; Margolis, H.; Bates, D.; Peters, J. (2004). The effect of air pollution on lung development from 10 to 18 years of age. N Engl J Med, 351: 1057

- 1067. http://nejm.highwire.org/cgi/content/abstract/351/11/1057. 056569

Gilboa, S. M.; Mendola, P.; Olshan, A. F.; Langlois, P. H.; Savitz, D. A.; Loomis, D.; Herring, A. H.; Fixler, D. E. (2005). Relation between ambient air quality and selected birth defects, seven county study, Texas, 1997

- 2000. Am J Epidemiol, 162: 238

- 252. 087892

Gilliland, F. D.; Berhane, K.; Rappaport, E. B.; Thoma s, D. C.; Avol, E.; Gauderman, W. J.; London, S. J.; Margolis, H. G.; McConnell, R.; Islam, K. T.; Peters, J. M. (2001). The effects of ambient air pollution on school absenteeism due to respiratory illnesses. Epidemiology, 12: 43 -54. 013232

Gilliland, F. D.; McConnell, R.; Peters, J.; Gong Jr, H. (1999). A Theoretical Basis for Investigating Ambient Air Pollution and Children's Respiratory Health. Environ Heal th Perspect, 107: 403 -407. 155792

Gilliland, F. D.; Rappaport, E. B.; Berhane, K.; Islam, T.; Dubeau, L.; Gauderman, W. J.; McConnell, R. (2002). Effects of glutathione S

- Transferase P1, M1, and T1 on acute respiratory illness in school children. Am J Respir Crit Care Med, 166: 346 -351. 090970

Giovannelli, L.; Pit ozzi, V.; Moretti, S.; Boddi, V.; Dolara, P. (2006). Seasonal variations of DNA damage in human lymphocytes: Correlation with different environmental variables. Mutat Res

- Fundam Mol Mech Mutagen, 593:

143- 152. http://dx.doi.org/10.1016/j.mrfmmm.2005.07.002 199894

Gonzalez

- Pina, R.; Escalante

- Membrillo, C.; Alfaro

- Rodriguez, A.; Gonzalez

- Maciel, A. (2008). Prenatal exposure to ozone disrupts cerebellar monoamine contents in newborn rats. Neurochem Res, 33: 912

- 918. http://dx.doi.org/10.1007/s11064-007- 9534 -3 475317

Gouveia, N.; Bremner, S. A.; Novaes, H. M. D. (200 4). Association between ambient air pollution and birth weight in Sao Paulo, Brazil. J Epidemiol Community Health, 58: 11 -17. 055613

Guevara- Guzmán, R.; Arria ga, V.; Kendrick, K. M.; Bernal, C.; Vega, X.; Mercado

- Gómez, O. F.; Rivas

- Arancibia, S. (2009). Estradiol prevents ozone

- induced increases in brain lipid peroxidation and impaired social recognition memory in female rats. Neuroscience, 159: 940

- 950. http: //dx.doi.org/10.1016/j.neuroscience.2009.01.047 596385

Ha, E.

- H.; Hong, Y.

- C.; Lee, B.

- E.; Woo, B.

- H.; Schwartz, J.; Christiani, D. C. (2001). Is air polluti on a risk factor for low birth weight in Seoul? Epidemiology, 12: 643 -648. 019390

Ha, E.

- H.; Lee, J.

- T.; Kim, H.; Hong, Y.

- C.; Lee (2003). Infant susceptibili ty of mortality to air pollution in Seoul, South Korea. Pediatrics, 111: 284

- 290. 042552

Hack, M.; Fanaroff, A. A. (1999). Outcomes of children of extremely l ow birth weight and gestational age in the 1990s. Early Hum Dev, 53: 193

- 218. http://dx.doi.org/10.1016/S0378

- 3782(98)00052 -8 625952

Hajat, S.; Armstrong, B. ; Wilkinson, P.; Busby, A.; Dolk, H. (2007). Outdoor air pollution and infant mortality: Analysis of daily time

- series data in 10 English cities. J Epidemiol Community Health, 61: 719 -722. 093276

March 20 11 7

- 66 DRAFT – DO NOT CITE OR QUOTEHanene, C.; Jihene, L.; Jame, A.; Kamel, H.; Agnes, H. (2007). Association of GST Genes Polymorphisms with Asthma in Tunisian Children. Mediators Inflamm, 19564: 6. 595428

Hansen, C. A.; Barnett, A. G.; Pritchard, G. (2008). The effect of ambient air pollution during early pregnancy on fetal ultrasonic measurements during mid

- pregnancy. Environ Health Perspect, 116: 362

- 36 9. 190273

Hansen, C.; Luben, T. J.; Sacks, J. D.; Olshan, A.; Jeffay, S.; Strader, L.; Perreault, S. D. (2010). The effect of ambient a ir pollution on sperm quality. Environ Health Perspect, 118: 203

- 209. http://dx.doi.org/10.1289/ehp.0901022 594438

Hansen, C.; Neller, A.; Williams, G.; Simpson, R. (2006). Matern al exposure to low levels of ambient air pollution and preterm birth in Brisbane, Australia. BJOG, 113: 935 -941. 089818

Hansen, C.; Neller, A.; Williams, G.; Simpson, R. (2007). Low levels of ambient air pollution during pregnancy and fetal growth among term neonates in Brisbane, Australia. Environ Res, 103: 383 -389. 090703

Harkema, J. R.; Plopper, C. G.; Hyde, D. M.; StGeorge, J. A.; Dungworth, D. L. (1987). Effects of an ambient level of ozone on primate nasal epithelial mucosubstances: quantitative histochemistry. Am J Pathol, 127: 90 -96. http://www.ncbi.nlm.nih. gov/pmc/articles/PMC1899606 /. 040816

Harkema, J. R.; Plopper, C. G.; Hyde, D. M.; StGeorge, J. A.; Wilson, D. W.; Dungworth, D. L. (1987). Response of the

mac aque nasal epithelium to ambient levels of ozone: a morphologic and morphometric study of the transitional and respiratory epithelium. Am J Pathol, 128: 29 -44. 041496

Harkema, J. R.; Plopper, C. G.; Hyde, D. M.; StGeorge, J. A.; Wilson, D. W.; Dungworth, D. L. (1993). Response of macaque bronchiolar epithelium to ambient concentrations of ozone. Am J Pathol, 143: 857

- 866. 039794

Haro, R.; Paz, C. (1993). Effects of ozone exposure during pregnancy on ontogeny of sleep in rats. Neurosci Lett, 164: 67

­ 70. http://dx.doi.org/10.1016/0304

- 3940(93)90859

- J 044194

Hassett, C.; Mustafa, M. G.; Coulson, W. F.; Elashoff, R. M. (1985). Murine lung carcinogenesis following exposure to ambient ozone concentrations. J Natl Cancer Inst, 75: 771

- 777. 040704

Hatch, G. E.; Harkema, J. R.; Plopper, C. G.; Harris, L. (1994). Ozone dosimetry studies in the nose and eye using oxygen

18. Inhal Toxicol, 6(suppl): 125

- 134. 076120

Himes, B. E.; Hunninghake, G. M.; Baurley, J. W.; Rafaels, N. M.; Sleiman, P.; Strachan, D. P.; Wilk, J. B.; Willis

- Owen, S.

A. G .; Klanderman, B.; Lasky

- Su, J.; Lazarus, R.; Murphy, A. J.; Soto

- Quiros, M. E.; Avila, L.; Beaty, T.; Mathias,

R. A.; Ruczinski, I.; Barnes, K. C.; Celedon, J. C.; Cookson, W. O. C.; Gauderman, W. J.; Gilliland, F. D.; Hakonarson, H.; Lange, C.; Moffatt, M. F.; O'Connor, G. T.; Raby, B. A.; Silverman, E. K.; Weiss, S. T. (2009).

Genome- wide Association Analysis Identifies PDE4D as an Asthma

- Susceptibility Gene. Am J Hum Genet, 84:

581- 593. http://dx.doi.org/10.1016/j.ajhg.2009.04.006 480112

Hollingsworth, J. W.; Free, M. E.; Li, Z.; Andrews, L. N.; Nakano, H.; Cook, D. N. (2010). Ozone activates pulmonary dendritic cells and promotes allergic sensitization through a Toll

- like receptor 4

- dependent mechanism. J Allergy Clin Immunol, 125: 1167 -1170. 635786

Horak, F. Jr; Studnicka, M.; Gartner, C.; Spengler, J. D.; T auber, E.; Urbanek, R.; Veiter, A.; Frischer, T. (2002). Particulate matter and lung function growth in children: a 3

- yr follow

- up study in Austrian schoolchildren. Eur Respir J, 19:

838-845. 034792

Huen, K.; Gunn, L.; Duramad, P.; Jeng, M.; Scalf, R.; Holland, N. (2006). Application of a geographic information system to explore associations between air pollution and micronucleus frequencies in African American children and adults. Environ Mol Mutagen, 47: 236

- 246. http://dx.doi.org/10.1002/em.20193 089035

Hutcheon, J. A.; Platt, R. W. (2008). The missing data problem in birth weight percentiles and thresholds for "small -forgestational

- age". Am J Epidemiol, 167: 786

- 792. 193795

Hwang, B. F.; Jaakkola, J. J. (2008). Ozo ne and other air pollutants and the risk of oral clefts. Environ Health Perspect,

116: 1411 -1415. 193794

Hwang, B.

- F.; Jaakkola, J. J. K.; Lee, Y.

- L.; Lin, Y
.- C.; Guo Y

- LL (2006). Relation between air pollution and allergic rhinitis in Taiwanese schoolchildren. Respir Res, 7: 23. 088971

Hwang, B.

- F.; Lee, Y.

- L.; L in, Y.

- C.; Jaakkola, J. J. K.; Guo, Y. L. (2005). Traffic related air pollution as a determinant of asthma among Taiwanese school children. Thorax, 60: 467 -473. 089454

March 20 11 7

- 67 DRAFT – DO NOT CITE OR QUOTEHyde, D. M.; Miller, L. A.; McDonald, R. J.; Stovall, M. Y.; Wong, V.; Pinkerton, K. E.; Wegner, C. D.; Rothlein, R.; Plopper, C. G. (1999). Neutrophils enhance clearance of necrotic epithelial cells in ozone

- induced lung injury in rhesus monkeys. Am J Physiol, 277: L1190 -L1198. 015124

Hyde, D. M.; Plopper, C. G.; Harkema, J. R.; St. George, J. A.; Tyler, W. S.; Dungworth, D. L. (1989). Ozone

- induced structural changes in monkey respiratory system. In T. Schneider; S. D. Lee; G. J. R. Wolters; L. D. Grant (Eds.), Atmospheric Ozone Research and its Policy Implications; Proceedings of the 3rd US

- Dutch International Symposium, Nijmegen, the Netherlands. M ay 9

- 13, 1988 (pp. 523

- 532). Amsterdam, the Netherlands: Elsevier

Science Publishers. 094057

Ihorst, G.; Frischer, T.; Horak, F.; Schumacher, M.; Kopp, M.; Fo rster, J.; Mattes, J.; Kuehr, J. (2004). Long

- and medium

term ozone effects on lung growth including a broad spectrum of exposure. Eur Respir J, 23: 292 -299. 055608

Islam, T.; Berhane, K.; McConnell, R.; Gauderman, W. J.; Avol, E.; Peters, J. M.; Gilliland, F. D. (2009). Glutathione -Stransferase (GST) P1, GSTM1, exercise, ozone and asthma incidence in school children. Thorax, 64: 197

- 202. http://dx.doi.org/10.1136/thx.2008.099366 196715

Islam, T.; Gauderman, W. J.; Berhane, K.; McConnell, R.; Avol, E.; Peters, J. M.; Gilliland, F. D. (2007). The relationship

betw een air pollution, lung function and asthma in adolescents. Thorax, 62: 957

- 963. 090697

Islam, T.; McConnell, R.; Gauderman, W. J.; Avol, E.; Peters, J. M.; G illiland, F. D. (2008). Ozone, oxidant defense genes and risk of asthma during adolescence. Am J Respir Crit Care Med, 177: 388

- 395. http://dx.doi.org/10.1164/rccm.200706- 863OC 097348

Jakab, G. J.; Bassett, D. J. P. (1990). Influenza virus infection, ozone exposure, and fibrogenesis. Am J Respir Crit Care Med, 141: 1307 -1315. 042196

Jalaludin, B.; Mannes, T.; Morgan, G.; Lincoln, D.; Sheppeard, V.; Corbett, S. (2007). Impact of ambient air pollution on gestational age is modified by season in Sydney, Australia. Environ Health, 6: 16. 156601

Jedlinska- Krakowska, M.; Bomba, G.; Jakubowski, K.; Rotkiewicz, T.; Jana, B.; Penkowskii, A. (2006). Impact of oxidative stress and supplementation with vitamins E and C on testes mo rphology in rats. J Reprod Dev, 52: 203

­ 209. 195640

Jerrett, M.; Burnett, R. T.; Pope, C. A.; Ito, K.; Thurston, G.; Krewski, D.; Shi, Y.; Calle, E.; Thun, M . (2009). Long

- term ozone exposure and mortality. N Engl J Med, 360: 1085

- 1095. http://dx.doi.org/10.1056/NEJMoa0803894 194160

Jiang L

- L; Zhang Y

- H; Song G

- X; Chen G

- H; Chen B

- H; Zhao N

- Q; Kan H

- D (2007). A time series analysis of outdoor air pollution and preterm birth in Shanghai, China. Biomed Environ Sci, 20: 426

- 431. 093029

Joad, J. P.; Bric, J. M.; Weir, A. J.; Putney, L.; Hyde, D. M.; Postlewait, E. M.; Plopper, C. G. (2000). Effect of respirato ry pattern on ozone injury to the airways of isolated rat lungs. Toxicol Appl Pharmacol, 169: 26 -32. 012984

Joad, J. P.; Kott, K. S.; Bric, J. M.; Peake, J. L.; Plopper, C. G.; Schelegle, E. S.; Gershwin, L. J.; Pinkerton, K. E. (200 6).

Structural and functional loc alization of airway effects from episodic exposure of infant monkeys to allergen and/or ozone. Toxicol Appl Pharmacol, 214: 237

- 243. http://dx.doi.org/10.1016/j.taap.2005.12.012 596390

Joad, J. P.; Kott, K. S.; Bric, J. M.; Schelegle, E. S.; Gershwin, L. J.; Plopper, C. G.; Peake, J. L.; Pinkerton, K. E. (200 8). The effects of inhaled corticosteroids on intrinsic responsiveness and histology of airways from in fant monkeys exposed to house dust mite allergen and ozone. Toxicol Appl Pharmacol, 226: 153 -160. http://dx.doi.org/10.1016/j.taap.2007.09.005 596391

Johnsto n, C.; Holm, B.; Gelein, R.; Finkelstein, J. (2006). Postnatal lung development: Immediate

- early gene responses post ozone and LPS exposure. Inhal Toxicol, 18: 875

- 883. http://dx.doi.org/10.1080/08958370600822466 097439

Kajekar, R.; Pieczarka, E. M.; Smiley

- Jewell, S. M.; Schelegle, E. S.; Fanucchi, M. V.; Plopper, C. G. (2007). Early postnatal exposure to allergen and ozone leads to hyperinnervation of the pul monary epithelium. Respir Physiol Neurobiol, 155: 55

- 63. http://dx.doi.org/10.1016/j.resp.2006.03.002 567661

Karr, C.; Lumley, T.; Schreuder, A.; Davis, R.; Larson, T.; Ritz, B.; Kaufman, J. (2007). Effects of subchronic and chronic exposure to ambient air pollutants on infant bronchiolitis. Am J Epidemiol, 165: 553

- 560. 090719

Kavlock, R. J.; Meyer, E.; Grabowski, C. T. (1980). Studies on the developmental toxicity of ozone: postnatal effects. Toxicol Lett, 5: 3 -9. 094043

Kavlock, R.; Daston, G.; Grabowski, C. T. (1979). Studies on the developmental toxicity of ozone I Prenatal effects. Toxicol Appl Pharmacol, 48: 19

- 28. http://dx.doi.org/10.1016/S0041

- 008X(79)80004 -6 039228

March 20 11 7

- 68 DRAFT – DO NOT CITE OR QUOTEKim, M. Y.; Cho, M. Y. (2009). Toxicity and carcinogenicity of ozone in combination with 4

- (N

- methyl -N- nitrosamino) -1­ (3

- pyridyl) -1- butanone and dibutyl phthalate in B6C3F1 mice for 16 and 32 weeks. Biomed Environ Sci, 22: 216

­ 222. 200775

Kim, M. Y.; Cho, M. Y. (2009). Tumorigenesis in B6C3F1 mice exposed to ozone in combination with 4

- (N

- methyl -Nnitrosami no)-1- (3

- pyridyl) -1- butanone and dietary dibutyl phthalate. Toxicol Ind Health, 25: 189

- 195. http://dx.doi.org/10.1177/0748233709106185 200773

Kinney, P. L.; Lippmann, M. (2000). Respiratory effects of seasonal exposures to ozone and particles. Arch Environ Occup Health, 55: 210 -216. 011913

Kleeberger, S. R.; Redd y, S. P.; Zhang, L.

- Y.; Cho, H.

- Y.; Jedlicka, A. E. (2001). Toll

- like receptor 4 mediates ozone

induced murine lung hyperpermeability via inducible nitric oxide synthase. Am J Physiol, 280: L326

- L333. 016163

Kodavanti, U. P.; Thomas, R.; Ledbetter, A. D.; Schladweiler, M. C.; Shannahan, J. H.; Wallenborn, J. G.; Lund, A. K.; Campen, M. J.; Butler, E. O.; Gottipolu, R. R.; Nyska, A.; Richards, J. E.; Andrews, D.; Jaskot, R. H.; McKee, J.; Kotha, S. R.; Patel, R. B.; Parianandi, N. L. (In Press). Vascular and cardiac impairments in rats Inhaling ozone and diesel exhaust particles. Environ Health Perspect. http://dx.doi.org/10.1289/ehp.1002386 666323

Krewski, D.; Jerrett, M.; Burnett, R. T.; Ma, R.; Hughes, E.; Shi, Y.; Turner, M. C.; Pope AC 3rd; Thurston, G.; Calle, E. E. ; Thun, M. J. (2009). Extended follow

- up and s patial analysis of the American Cancer Society study linking particulate air pollution and mortality (Report No. Report Nr. 140). Cambridge, MA: Health Effects Institute.

191193

Kuo, H. W.; Lai, J. S.; Lee, M. C.; Tai, R. C.; Lee, M. C. (2002). Respiratory effects of air pollutants among asthmatics in central Taiwan. Arch Environ Occup Health, 57: 194

- 200. 036310

Lampl, M.; Jeanty, P. (2003). Timing is everything: A reconsideration of fetal growth velocity patterns identifies the importance of individual and sex differences. Am J Hum Biol, 15: 667

- 680. http://dx.do i.org/10.1002/ajhb.10204

625252

Larson, S. D.; Schelegle, E. S.; Walby, W. F.; Gershwin, L. J.; Fanuccihi, M. V.; Evans, M. J.; Joad, J. P.; Tarkington, B. K .; Hyde, D. M.; Plopper, C. G. (2004). Postnatal remodeling of the neural components of the epithelial

- mesenchymal trophic unit in the proximal airways of infant rhesus monkeys exposed to ozone and allergen. Toxicol Appl Pharmacol, 194: 211

- 220. 057062

Last, J. A.; Warren, D. L.; Pecquet

- Goad, E.; Witschi, H. (1987). Modification by ozone of lung tumor development in mice. J Natl Cancer Inst, 78: 149 -154. 040830

Latzin, P.; Röösli, M.; Huss, A.; Kuehni, C. E.; Frey, U. (2009). Air pollution during pregnancy and lung function in newborns: A birth cohort study. Eur Respir J, 33: 594 -603. 195721

Lee, C.; Watt, K. C.; Chang, A. M.; Plopper, C. G.; Buckpitt, A. R.; Pinkerton, K. E. (1998). Site

- selective differences in
cytoc hrome P450 isoform activities: comparison of expression in rat and rhesus monkey lung and induction in rats. Drug Metab Dispos, 26: 396

- 400. 054473

Lee, S. J. ; Hajat, S.; Steer, P. J.; Filippi, V. (2008). A time

- series analysis of any short

- term effects of meteorological and air pollution factors on preterm births in London, UK. Environ Res, 106: 185 -194. 195720

Lee, Y. L.; McConnell, R.; Berhane, K.; Gilliland, F. D. (2009). Ambient ozone modifies the effect of tumor necrosis factor

G- 308A on bronchitic symptoms among children with asthma. Allergy, 64: 1342 -1348. http://dx.doi.org/10.1111/j.1398-9995.2009.02014.x 199915

Lee, Y.

- L.; Lin, Y.

- C.; Lee, Y.

- C.; Wang, J.

- Y.; Hsiue, T.

- R.; Guo, Y. L. (2004). Glutathione S

- tra nsferase P1 gene polymorphism and air pollution as interactive risk factors for childhood asthma. Clin Exp Allergy, 34: 1707 -1713.

090971

Legro, R. S.; Sauer, M. V.; Mottla, G. L.; Richter, K. S.; Li, X.; Dodson, W. C.; Liao, D. (2010). Effect of air quality on assisted human reproduction [in press]. Hum Reprod, TBD: TBD. http://dx.doi.org/10.1093/humrep/deq021 597377

Li, H.; Romieu, I.; Sienra

- Monge, J. J.; Ramirez

- Aguilar, M.; Estela del Rio

- Navarro, B.; Kistner, E. O.; Gjessing, H. K.;Irma del Carmen, L. S.; Chiu, G. Y.; London, S. J. (2006). Genetic polymorphis ms in arginase I and II and childhood asthma and atopy. J Allergy Clin Immunol, 117: 119 –126. 596447

March 20 11 7

- 69 DRAFT – DO NOT CITE OR QUOTELi, Y. F.; Gauderman, W. J.; Conti, D. V.; Lin, P. C.; A vol, E.; Gilliland, F. D. (2008). Glutathione S

- Transferase P1, Maternal Smoking, and Asthma in Children: A Haplotype

- Based Analysis. Environ Health Perspect, 116: 409 -415.

596449

Lin, C. A.; Pereira, L. A. A.; Nishioka, D. C.; Conceicao, G. M. S.; Graga, A. L. F.; Saldiva, P. H. N. (2004). Air pollution and neonatal deaths in Sao Paulo, Brazil. Braz J Med Biol Res, 37: 765

- 770. 095787

Lin, C.

- M.; Li, C.

- Y.; Yang, G.

- Y.; Mao, I.

- F. (2004). Association between maternal exposure to elevated ambient sulfur dioxide during pregnancy and term low birth weight. Environ Re s, 96: 41 -50. 089503

Lin, S.; Fitzgerald, E.; Hwang, S. A.; Munsie, J. P.; Stark, A. (1999). Asthma hospitalization rates and socioeconomic status

in New York State (1987

- 1993). J Asthma, 36: 239

- 251. http://informahealthcare.com/doi/abs/10.3109/02770909909075407

. 377917

Lin, S.; Liu, X.; Le, L. H.; Hwang, S. A. ( 2008). Chronic exposure to ambient ozone and asthma hospital admissions among children. Environ Health Perspect, 116: 1725

- 1730. http://dx.doi.org/10.1289/ehp.11184 196680

Linn, W. S.; Rappaport, E. B.; Berhane, K. T.; Bastain, T. M.; Avol, E. L.; Gilliland, F. D. (2009). Exhaled nitric oxide in a population- based study of southernCalifornia schoolchildren. Respir Res, 10: 28. 597363

Lipfert, F. W.; Baty, J. D.; Miller, J. P.; Wyzga, R. E. (2006). PM2.5 constituents and related air quality variables as predictors of survival in a cohort of U.S. military veterans. Inhal Toxicol, 18: 645

- 657. 088756

Lipfert, F. W.; Perry, H. M. Jr; Miller, J. P.; Baty, J. D.; Wyzga, R. E.; Carmody, S. E. (2000). The Washington University

EPRI veterans' cohort mortality study: Preliminary results. Inhal Toxicol, 4: 41 -73. 004087

Lipfert, F. W.; Perry, H. M. Jr; Miller, J. P.; Baty, J. D.; Wyzga , R. E.; Carmody, S. E. (2003). Air pollution, blood pressure, and their long

- term associations with mortality. Inhal Toxicol, 15: 493

- 512. 052250

Lipfert, F. W.; Wyzga, R. E.; Baty, J. D.; Miller, J. P. (2006). Traffic density as a surrogate measure of environmental exposures in studies of air pollution health effects: Long

- term mortality in a cohort of US veterans. Atmos Environ,

40: 154-169. 088218

Liu, S.; Krewski, D.; Shi, Y.; Chen, Y.; Burnett, R. (2007). Association between maternal exposure to ambient air pollutants during pregnancy and fetal growth restrict ion. J Expo Sci Environ Epidemiol, 17: 426

- 432. 090429

Loomis, D.; Castillejos, M.; Gold, D. R.; McDonnell, W.; Borja

- Aburto, V. H. (1999). Air pollution and infant mortality in Mexico City. Epidemiology, 10: 118 -123. 087288

López, I.; Sánchez, I.; Bizarro, P.; Acevedo, S.; Ustarroz, M.; Fortoul, T. (2008). Ultrast ructural alterations during embryonic rats' lung development caused by ozone. J Electron Microsc (Tokyo), 57: 19 -23. http://dx.doi.org/10.1093/jmicro/dfm033 1 97786

Maisonet, M.; Correa, A.; Misra, D.; Jaakkola, J. J. (2004). A review of the literature on the effects of ambient air polluti on on fetal growth. Environ Res, 95: 106 -115. 156725

Mannes, T.; Jalaludin, B.; Morgan, G.; Lincoln, D.; Sheppeard, V.; Corbett, S. (2005). Impact of ambient air pollution on birth weight in Sydney, Australia. Occup Environ Med, 62: 524 -530. 087895

Mariassy, A. T.; Abraham, W. M.; Phipps, R. J.; Sielczak, M. W.; Wanner, A. (1990). Effect of ozone on the postnatal development of lamb mucociliary apparatus. J Appl Physiol, 68: 2504 -2510. 042311

Mariassy, A. T.; Sielczak, M. W.; McCray, M. N.; Abraham, W. M.; Wanner, A. (1989). Effects of ozone on lamb tracheal mucosa: Quantitative glycoconjug ate histochemistry. Am J Pathol, 135: 871 -879. 042246

Marshall, E.; Harris, G.; Wartenberg, D. (2010). Oral cleft defects and maternal exposure to ambient air pollutants in New Jersey. Birth Defects Res A Clin Mol Teratol, 88: 205

- 215. http://dx.doi.org/10.1002/bdra.20650 597374

Martinez, F. D.; Wright, A. L.; Tau ssig, L. M.; Holberg, C. J.; Halonen, M.; Morgan, W. J.; Group Health Medical Associates (1995). Asthma and wheezing in the first six years of life. N Engl J Med, 332: 133

- 138. 046150

McConnell, R.; Berhane, K.; Gilliland, F.; London, S. J.; Islam, T.; Gauderman, W. J.; Avol, E.; Margolis, H. G.; Peters, J.

M. (2002). Asthma in exercising children exposed to ozone: A cohort study. Lancet, 359: 386 -391. 023150

McConnell, R.; Islam, T.; Shankardass, K.; Jerrett, M.; Lurmann, F.; Gilliland, F.; Gauderman, J.; Avol, E.; Kuenzli, N.; Yao, L.; Peters, J.; Berhane, K. (201 0). Childhood incident asthma and traffic

- related air pollution at home and school. Environ Health Perspect, TBD: TBD. http://dx.doi.org/10.1289/ehp.0901232 6 25501

March 20 11 7

- 70 DRAFT – DO NOT CITE OR QUOTEMeng, Y. Y.; Rull, R. P.; Wilhelm, M.; Lombardi, C.; Balmes, J.; Ritz, B. (2010). Outdoor air pollution and uncontrolled asthma in the San Joaquin Valley, California. J Epidemiol Community Health, 64: 142

- 147. http://dx.doi.org/10.1136/jech.2008.083576 594252

Meng, Y. Y.; Wilhelm, M.; Rull, R. P.; English, P.; Ritz, B. (2007). Traffic and outdoor air pollution levels near residences and poorly controlle d asthma in adults. Ann Allergy Asthma Immunol, 98: 455

- 463. 093275

Miller, L. A.; Gerriets, J. E.; Tyler, N. K.; Abel, K.; Schelegle, E. S.; Plopper, C. G.; Hyde, D. M. (2009). Ozone and allergen exposure during postnatal development alters the frequency and airway distribution of CD25+ cells in infant rhesus monkeys. Toxicol Appl Pharmacol, 236: 39

- 48. http://dx.doi.org/10.1016/j.taap.2008.12.031 596406

Moffatt, R. K.; Hyde, D. M.; Plopper, C. G.; Tyler, W. S.; Putney, L. F. (1987). Ozone

- induced adaptive and reactive cellular changes in respiratory bronchioles of Bonnet monkeys. Exp Lung Res, 12: 57 -74. 040841

Monchaux, G.; Morlier, J. P.; Morin, M.; Rochefort, P.; Maximilien, R.; Tredaniel, J. (1996). Co -carcinogeni c effects in rats of combined exposure to radon and ozone. Environ Int, 221: S909

- S915. 086197

Moore, K.; Neugebauer, R.; Lurmann, F.; Hall, J.; Brajer, V.; A lcorn, S.; Tager, I. (2008). Ambient ozone concentrations cause increased hospitalizations for asthma in children: An 18

- year study in Southern California. Environ Health Perspect, 116: 1063

- 1070. http://dx.doi.org/10.1289/ehp.10497 196685

Morello- Frosch, R.; Jesdale, B. M.; Sadd, J. L.; Pastor, M. (2010). Ambient air pollution exposure and full

- term birth weight in California. Environ Health, 9: 44. http://dx .doi.org/10.1186/1476- 069X -9-44 670076

Morris, C. R.; Poljakovic, M.; Lavrisha, L.; Machado, L.; Kuypers, F. A.; Morris, S. M. Jr (2004). Decreased arginine bioavailability and increased serum arginase activity in asthma. Am J Respir Crit Care Med, 170: 148

- 153. http://dx.doi.org/10.1164/rccm.200309- 1304OC 674145

Mortimer, K.; Neugebauer, R.; Lurmann, F.; Alcorn, S.; Balmes, J.; Tager, I. (2008). Air pollution and pulmonary function in asthmatic children: Effects of prenatal and lifetime exposures. Epidemiology, 19: 550 -557. http://dx.doi.org/10.1097/EDE.0b013e31816a9dcb 122163

Mortimer, K.; Neugebauer, R.; Lurmann, F.; Alcorn, S.; Balmes, J.; Tager, I. (2008). Early

- Lifetime exposure to air pollution and allergic sen sitization in children with asthma. J Asthma, 45: 874

- 881. http://dx.doi.org/10.1080/02770900802195722 187280

NTP (1994). Toxicology and carcinogenesis: studies of ozone (CAS No 10028

- 15

- 6) and ozone/NNK (CAS No 10028

- 15

­ 6/64091 -91- 4) in F344/N rats and B6C3F1 mice (Report No. Technical Report No. 440). Research Triangle Park, NC: National Toxicology Program. http://ntp.niehs.nih.gov/index.cfm?objectid=070A0EB D- 081E

- B501

­ E38F640803C3542C. 011143

Oryszczyn, M. P.; Bouzigon, E.; Maccario, J.; Siroux, V.; Nadif, R.; Wright, A.; Kauffmann, F. (2007). Interrelationships of quantitative asthma

- related phenotypes in the Epidemiological Study on the Genetics and Environment of Asthma, Bronchial Hyperresponsiveness, and Atopy. J Allergy Clin Immunol, 119: 57 -63. 596460

Palli, D.; Sera, F.; Giovannelli, L.; Masala, G.; Grechi, D.; Bendinelli, B.; Caini, S.; Dolara, P.; Saieva, C. (2009). Environmental ozone exposure and oxidative DNA damage in adult residents of Florence, Italy. Environ Pollut,

157: 1521?1525. http://dx.doi.org/10.1016/j.envpol.2008.09.011 196688

Parker, J. D.; Akinbami, L. J.; Woodruff, T. J. (2009). Air pollution a nd childhood respiratory allergies in the United States. Environ Health Perspect, 117: 140

- 147. http://dx.doi.org/10.1289/ehp.11497 192359

Paz, C.; Bazan

- Per kins, B. (1992). Sleep

- wake disorganization in cats exposed to ozone. Neurosci Lett, 140: 270

- 272.

036436

Paz, C.; Huitron

- Resendiz, S. (1996). The effects of ozone exposure on the sleep

- wake cycle and serotonin contents in the pons of the rat. Neurosci Lett, 204: 49 -52. 082684

Peluso M Hainaut, P.; Airoldi, L.; Autrup, H.; Dunning, A.; Garte, S.; Gormally, E.; Malaveille, C.; Matullo, G.; Munniaa, A.; Riboli, E.; Vineis P EPIC investigators (2005). Methodology of laboratory measurements in prospective studies

on gene

- environment interactions: The experience of GenAir. DNA Repair, 574: 92 -104. 089057

Penard

- Morand, C.; Charpin, D.; Raherison, C.; Kopferschmitt, C.; Caillaud, D.; Lavaud, F.; Annesi

- Maesano, I. (2005). Long- term exposure to background air pollution related to respiratory and allergic health in schoolchildren. Clin Exp Allergy, 35: 1279 -1287. 087951

March 20 11 7

- 71 DRAFT – DO NOT CITE OR QUOTEPereira, F. A. C.; De Assuncao, J. V.; Saldiva, P. H. N.; Pereira, L. A. A.; Mirra, A. P.; Braga, A. L. F. (2005). Influence of air pollution on the incidence of respiratory tract neoplasm. J Air Waste Manag Assoc, 55: 83 -87. http://www.ncbi.nlm.nih.gov/pubmed/=15704542

. 073851

Pereira, L. A. A.; Loomis, D.; Conceicao, G. M. S.; Braga, A. L. F.; Arcas, R. M.; Kishi, H. S.; Singer, J. M.; Bohm, G. M.; Saldiva, P. H. N. (199 8). Association between air pollution and intrauterine mortality in Sao Paulo, Brazil. Environ Health Perspect, 106: 325

- 329. 007264

Perepu, R. S.; Garcia, C.; Dostal, D.; Sethi, R. (2010). Enhanced death signaling in ozone

- exposed ischemic -reperfused hearts. Mol Cell Biochem, 336: 55

- 64. http://dx.doi.org/10.1007/s11010 -009- 0265 -4 385020

Peters, J. M.; Avol, E.; Gauderman, W. J.; Linn, W. S.; Navidi, W.; London, S. J.; Margolis, H.; Rappaport, E.; Vora, H.; Gong, H. Jr; Thomas, D. C. (1999). A study of twelve southern California communities with differing levels a nd types of air pollution II Effects on pulmonary function. Am J Respir Crit Care Med, 159: 768 -775. 087237

Peters, J. M.; Avol, E.; Navidi, W.; London, S. J.; Gauderman, W. J.; Lurmann, F.; Linn, W. S.; Margolis, H.; Rappaport, E.; Gong, H. Jr; Thomas, D. C. (1999). A study of twelve southern California communities with differing levels and types of air pollution I Prevalence of respiratory morbidity. Am J Re spir Crit Care Med, 159: 760

- 767. 087243

Petruzzi, S.; De Acetis, L.; Chiarotti, F.; Sorace, A.; Alleva, E. (1999). Limited changes in handedness and morphine reactivity in CD

- 1 mice after pre

- and postnatal ozone exposure. Acta Neurobiol Exp (Wars), 59: 115

- 122. 012066

Petruzzi, S.; Fiore, M.; Dell'Omo, G.; Bignam i, G.; Alleva, E. (1995). Medium and long

- term behavioral effects in mice of extended gestational exposure to ozone. Neurotoxicol Teratol, 17: 463

- 470. 077448

Plopper, C. G.; Harkema, J. R.; Last, J. A.; Pinkerton, K. E.; Tyler, W. S.; StGeorge, J. A.; Wong, V. J.; Nishio, S. J.; Wei r, A. S.; Dungworth, D. L.; Barry, B. E.; Hyde, D. M. (1991). The respiratory system of nonhuman primates responds

more to ambient concentrations of ozone than does that of rats. In Tropospheric Ozone and the Environment (pp.

137- 150). Pittsburgh, PA: Air and Waste Management Association. 042617

Plopper, C. G.; Hatch, G. E.; Wong, V.; Duan, X.; Weir, A. J.; Tarkington, B. K.; Devlin, R. B.; Becker, S.; Buckpitt, A. R. (1998). Relationship of inhaled ozone concentration to acute tracheobronchial epithelial injury, site

- specific ozone

dose and glutathione depletion in rhesus monkeys. Am J Respir Cell Mol Biol, 19: 387

- 399. 087203

Plopper, C. G.; Schelegle, E. S. (1997). Report of on

- going and fu ture research on ozone at the Davis campus of the University of California. Washington, DC: U.S. Environmental Protection Agency. 656713

Plopper, C. G.; Smil ey

- Jewell, S. M.; Miller, L. A.; Fanucchi, M. V.; Evans, M. J.; Buckpitt, A. R.; Avdalovic, M.; Gershwin, L. J.; Joad, J. P.; Kajekar, R.; Larson, S.; Pinkerton, K. E.; Van Winkle, L. S.; Schelegle, E. S.; Pieczarka,

E. M.; Wu, R.; Hyde, D. M. (2007). Asth ma/allergic airways disease: Does postnatal exposure to environmental toxicants promote airway pathobiology? Toxicol Pathol, 35: 97

- 110. http://dx.doi.org/10.1080/01926230601132030

596412

Pope CA 3rd; Burnett, R. T.; Thun, M. J.; Calle, E. E.; Krewski, D.; Ito, K.; Thurston, G. D. (2002). Lung cancer, cardiopulmonary mortality, and long

- term exposure to fine particulate air pollution. JAMA, 287: 1132 -1141.

024689

Qian, Z.; Liao, D.; Lin, H.

- M.; Whitsel, E. A.; Rose, K. M.; Duan, Y. (2005). Lung function and long

- term exposure to air pollutants in middle

- aged Americ an adults. Arch Environ Occup Health, 60: 156

- 163. 093283

Rage, E.; Jacquemin, B.; Nadif, R.; Oryszczyn, M. P.; Siroux, V.; Aguilera, I.; Kauffmann, F.; Kunzl i, N. (2009). Total serum IgE levels are associated with ambient ozone concentration in asthmatic adults. Allergy, 64: 40 -46. 196719

Rage, E.; Siroux, V.; Ku nzli, N.; Pin, I.; Kauffmann, F. (2009). Air pollution and asthma severity in adults. Occup Environ Med, 66: 182

- 188. http://dx.doi.org/10.1136/oem.2007.038349 196720

Ramadour, M.; Burel, C.; Lanteaume, A.; Vervloet, D.; Charpin, D.; Brisse, F.; Dutau, H.; Charpin, D. (2000). Prevalence of asthma and rhinitis in relation to long

- term exposure to gaseous air pollutants. Allergy, 55: 1163 -1169. 013259

Renzetti, G.; Silvestre, G.; D'Amario, C.; Bottini, E.; Gloria

- Bottini, F.; Bottini, N.; Auais, A.; Perez, M. K.; Piedimonte, G. (2009). Less air pollution leads to rapid reduction of airway inflammation and improved airway function in asthmatic children. Pediatrics, 123: 1051

- 1058. http://dx.doi.org/10.1542/peds.2008 -1153 1 99834

Ritz, B.; Wilhelm, M. (2008). Ambient air pollution and adverse birth outcomes: Methodologic issues in an emerging field. Basic Appl Ecol, 102: 182

- 190. 156914

March 20 11 7

- 72 DRAFT – DO NOT CITE OR QUOTERitz, B.; Wilhelm, M.; Hoggatt, K. J.; Ghosh, J. K. (2007). Ambient air pollution and preterm birth in the environment and pregnancy outcomes study at the University of California, Los Angeles. Am J Epidemiol, 166: 1045 -1052. 096146

Ritz, B.; Wilhelm, M.; Zhao, Y. (2006). Air pollution and infant death in southern California, 1989

- 2000. Pediatrics, 118:

493-502. 089819

Ritz, B.; Yu, F. (1999). The effect of ambient carbon monoxide on low birth weight among children born in southern California between 1989 and 1993. Environ Health Perspect, 107: 17 -25. 086976

Ritz, B.; Yu, F.; Chapa, G.; Fruin, S. (2000). Effect of air pollution on preterm birth among children born in Southern California between 1989 and 1993. Epidemiology, 11: 502 -511. 012068

Ritz, B.; Yu, F.; Fruin, S.; Chapa, G.; Shaw, G. M.; Harris, J. A. (2002). Ambient air pollution and risk of birth defects in Southern California. Am J Epidemiol, 155: 17 -25. 023227

Rivas

- Arancibia, S.; Guevara

- Guzmán, R.; López

- Vidal, Y.; Rodríguez

- Martínez, E.; Gomes, M. Z.; Angoa

- Pérez, M.;

Raisman

- Vozari, R. (2010). Oxidative stress caused by ozone exposure induces loss of brain repair in the hippocampus of adult rats. Toxicol Sci, 113: 187

- 197. http://dx.doi.org/10.1093/toxsci/kfp252 201544

Rivas

- Manzano, P.; Paz, C. (1999). Cerebellar morphological alterations in rats induced by prenatal ozone exposure. Neurosci Lett, 276: 37 -40. 012072

Rojas- Martinez, R.; Perez

- Padilla, R.; Olaiz

- Fernandez, G.; Mendoza

- Alvarado, L.; Moreno

- Macias, H.; Fortoul, T.; McDonnell, W.; Loomis, D.; Romieu, I. (2007). Lung function growth in children with long

- term exposure to air pollutants in Mexico City. Am J Respir Crit Care Med, 176: 377 -384. 091064

Romero, R.; Espinoza, J.; Kusanovic, J. P.; Gotsch, F.; Hassan, S.; Erez, O.; Chaiworapongsa, T.; Mazor, M. (2006). The preterm parturition syndrome. BJOG, 113: 17

- 42. http://dx.doi.org/10.1111/j.1471 -0528.2006.01120.x 625253

Romero

- Velazquez, R. M.; Alfaro

- Rodriguez, A.; Gonzalez

- Pina, R.; Gonzalez

- Maciel, A. (2002). Effect of ozone prenatal exposure on postnatal development of cerebellum. Proc West Pharmacol Soc, 45: 65 -67. 035575

Romieu, I.; Ramirez

- Aguilar, M.; Moreno

- Macias, H.; Barraza

- Villarreal, A.; Miller, P.; Hernandez

- Cadena, L.; Carbajal

Arroyo, L. A.; Hernandez

- Avila, M. (2004). Infant mortality and air pollution : Modifying effect by social class. J Occup Environ Hyg, 46: 1210 -1216. 093074

Romieu, I.; Sienra

- Monge, J. J.; Ramirez

- Aguilar, M.; Moreno

- Macias, H.; Reyes

- Ruiz, N. I.; Estela del Rio

- Navarro, B.;

Hernandez- Avila, M.; London, S. J. (2004). Genetic polymorphism of GSTM1 and antioxidant supplementation influence lung function in relation to ozone exposure in asthmatic children in Mexico City. Thorax, 59: 8 -10.

056796

Rubes, J.; Selevan, S. G.; Evenson, D. P.; Zudova, D.; Vozdova, M.; Zudova, Z.; Robbins, W. A.; Perreault, S. D. (2005). Episodic air pollution is asso ciated with increased DNA fragmentation in human sperm without other changes in semen quality. Hum Reprod, 20: 2776 -2783. 078091

Rubio, C.; Paz, C. (2003). In domethacin reverts sleep disorders produced by ozone exposure in rats. Toxicology, 191: 89

­ 96. http://dx.doi.org/10.1016/S0300

- 483X(03)00245 -2 053541

Salam, M . T.; Islam, T.; Gauderman, W. J.; Gilliland, F. D. (2009). Roles of arginase variants, atopy, and ozone in childhood asthma. J Allergy Clin Immunol, 123: 596

- 602. http://dx.doi.org/10.1016/j.jaci.2008.12.020 596644

Salam, M. T.; Millstein, J.; Li, Y.

- F.; Lurmann, F. W.; Margolis, H. G.; Gilliland, F. D. (2005). Birth outcomes and prenatal exposure to ozone, carbon monoxide, and particulate matter: Results fro m the Children's Health Study. Environ Health Perspect, 113: 1638 -1644. 087885

Santucci, D.; Sorace, A.; Francia, N.; Aloe, L.; Alleva, E. (2006). Prolonged p renatal exposure to low

- level ozone affects aggressive behaviour as well as NGF and BDNF levels in the central nervous system of CD

- 1 mice. Behav Brain Res, 166: 124

- 130. http://dx.doi.org/10.1016/j.bbr.2005.07.032 596414

Schatz, M.; Zeiger, R. S.; Hoffman, C. P. (1990). Intrauterine growth is related to gestational pulmonary function in pregnant asthmatic women. Kaiser

- Permanente Asthma and Pregnancy Study Gr oup. Chest, 98: 389

- 392. 156073

Schelegle, E. S.; Miller, L. A.; Gershwin, L. J.; Fanucchi, M. V.; Van Winkle, L. S.; Gerriets, J. E.; Walby, W. F.; Mitchell

, V.; Tarkington, B. K.; Wong, V. J.; Baker, G. L.; Pantle, L. M.; Joad, J. P.; Pinkerton, K. E.; Wu, R.; Evans, M. J.; Hyde, D. M.; Plopper, C. G. (2003). Repeated episodes of ozone inhalation amplifies the effects of allergen sensitization and inhalation on airway immune and structural development in Rhesus monkeys. Toxicol Appl Pharmacol, 191: 74 -85. 053778

March 20 11 7

- 73 DRAFT – DO NOT CITE OR QUOTESchmelzer, K. R.; Wheelock, A. M.; Dettmer, K.; Mor in, D.; Hammock, B. D. (2006). The role of inflammatory mediators in the synergistic toxicity of ozone and 1

- nitronaphthalene in rat airways. Environ Health Perspect, 114: 1354 -1360.

112994

Schöpke, R.; Wolfer, D. P.; Lipp, H. P.; Leisinger

- Trigona, M. C. (1991). Swimming navigation and structural variations of the infrapyramidal mossy fibers in the hippocampus of the mouse. Hippocampus, 1: 315 -328. http://dx.doi.org/10.1002/hipo.450010322 684254

Selevan, S. G.; Borkovec, L.; Slott, V. L.; Zudova, Z.; Rubes, J.; Evenson, D. P.; Perreault, S. D. (2000). Semen quali ty and reproductive health of young Czech men exposed to seasonal air pollution. Environ Health Perspect, 108: 887 -894.

012578

Slama, R.; Darrow, L.; Parker, J.; Woodruff, T. J.; Strickland, M.; Nieuwenhuijsen, M.; Glinianaia, S.; Hoggatt, K. J.; Kannan, S.; Hurley, F.; Kalinka, J.; Sram, R.; Brauer, M.; Wilhelm, M.; Heinrich, J.; Ritz, B. (2008). Meeting report: Atmospheric pollution and human reproduction. En viron Health Perspect, 116: 791 -798. 156985

Sokol, R. Z.; Kraft, P.; Fowler, I. M.; Mamet, R.; Kim, E.; Berhane, K. T. (2006). Exposure to environmental ozon e alters semen quality. Environ Health Perspect, 114: 360

- 365. 098539

Son, J. Y.; Cho, Y. S.; Lee, J. T. (2008). Effects of air pollution on postneonatal infa nt mortality among firstborn infants in Seoul, Korea: Case

- crossover and time

- series analyses. Arch Environ Occup Health, 63: 108 -13. 190323

Sousa, S. I. V.; Alvim

- Ferraz, M. C. M.; Martins, F. G.; Pereira, M. C. (2009). Ozone exposure and its influence on the worsening of childhood asthma. Allergy, 64: 1046

- 1055. http://dx.doi.org/10.1111/j.1398 -9995.2009.01946.x

619956

Sousa, S. I. V.; Pereira, M. M. C.; Martins, F. G.; Alvim

- Ferraz, C. M. (2008). Identification of regions with high ozone concentrations aiming the impact assessment on childhood asthma. Hum Ecol Risk Assess, 14: 610 -622. http://dx.doi.org/10.1080/10807030802074147 619959

Sousa, S. I.; Ferraz, C.; Alvim

- Ferraz, M. C.; Martins, F. G.; Vaz, L. G.; Perei ra, M. C. (2011). Spirometric tests to assess the prevalence of childhood asthma at Portuguese rural areas: Influence of exposure to high ozone levels. Environ Int, 37: 474

- 478. http://dx.doi.org/10.1016/j.envint.2010.11.014 676712

Sram, R. J.; Binkova, B.; Dejmek, J.; Bobak, M. (2005). Ambient air pollution and pregnancy outcomes: a review of the literature. Environ Health Perspect, 113: 375

- 382. 087442

Sram, R. J.; Binkova, B.; Rossner, P.; Rubes, J.; Topinka, J.; Dejmek, J. (1999). Adverse reproductive outcomes from exposure to environmental mutagens. Mutat Res

- Fundam Mol Mech Mutagen, 428: 203 -215. http://dx.doi.org/10.1016/S1383-5742(99)00048-4 078127

Stedman, J. R.; Kent, A. J. (2008). An analysis of the spatial p atterns of human health related surface ozone metrics across the UK in 1995, 2003 and 2005. Atmos Environ, 42: 1702 -1716. 110057

Stokinger, H. E. (1962). Effects of air pollution in animals. In AC Stern (Ed.), Air pollution (pp. 282

- 334). New York, NY:

Academic Press. 015101

Strickland, M. J.; Klein, M.; Correa, A .; Reller, M. D.; Mahle, W. T.; Riehle

- Colarusso, T. J.; Botto, L. D.; Flanders, W. D.; Mulholland, J. A.; Siffel, C.; Marcus, M.; Tolbert, P. E. (2009). Ambient air pollution and cardiovascular malformations in Atlanta, Georgia, 1986

- 2003. Am J Epidemiol, 169: 1004 -14. 190324

Tager, I. B.; Balmes, J.; Lurmann, F.; Ngo, L.; Alcorn, S.; Kunzli, N. (2005). Chronic exposure to ambient ozone and lung function in y oung adults. Epidemiology, 16: 751

- 759. http://dx.doi.org/10.1097/01.ede.0000183166.68809.b0

087538

Tamer, L.; Caliko?lu, M.; Ates, N. A.; Yildirim, H.; Ercan , B.; Saritas, E.; Unlü, A.; Atik, U. (2004). Glutathione -Stransferase gene polymorphisms (GSTT1, GSTM1, GSTP1) as increased risk factors for asthma. Respirology, 9:

493-498. 199914

Tovalin, H.; Valverde, M.; Morandi, M. T.; Blanco, S.; Whitehead, L.; Rojas, E. (2006). DNA damage in outdoor workers occupationally exposed to environmental air pollutants. Occup Environ Med, 63: 230

- 236. 091322

Tran, M. U.; Weir, A. J.; Fanucchi, M. V.; Rodriguez, A. E.; Pantle, L. M.; Smiley

- Jewell, S. M.; Van Winkle, L. S.; Evans,

M. J.; Miller, L. A.; Schelegle, E. S.; Gershwin, L. J.; Hyde, D. M.; Plopper, C. G. (2004). Smooth muscle hypertrophy in distal airways of sensitized infant rhesus monkeys exposed to house dust mite allergen. Clin Exp Allergy, 34: 1627

- 1633. http://dx.doi.org/10.1111/j.1365 -2222.2004.02057.x 628626

March 20 11 7

- 74 DRAFT – DO NOT CITE OR QUOTETsai, S.

- S.; Chen, C.

- C.; Hsieh, H.

- J.; Chang, C.

- C.; Yang, C.

- Y. (2006). Air pollution and postneonatal mortality in a tropical city: Kaohsiung, Taiwan. Inhal Tox icol, 18: 185 -189. 090709

Tucker, J. D.; Auletta, A.; Cimino, M. C.; Dearfield, K. L.; Jacobson

- Kram, D.; Tice, R. R.; Carrano, A. V. (1993). Sister

chromatid exchange: second report of the GENE

- TOX program. DNA Repair, 297: 101 -180. http://dx.doi.org/10.1016/0165- 1110(93)90001 -4 056406

U.S. EPA (1996). Air quality criteria for ozone and related photochemical oxidants (Report No. EPA/600/P

- 93/004AF). Research Triangle Park, NC: U.S. Environmental Protection Agency, Office of Research and Development. (NTIS No. PB94173127) 017831

U.S. EPA (2006). Air quality criteria for ozone and related photochemical oxidants (Report No. EPA/600/R -05/004AF). Research Triangle Park, NC: U.S. Environmental Protection Agency, Office of Research and Development. http://cfpub.epa.gov/ncea/cfm/recordisplay.cfm?deid=149923

. 088089

Vancza, E. M.; Galdanes, K.; Gunnison, A.; Hatch, G.; Gordon, T. (20 09). Age, strain, and gender as factors for increased sensitivity of the mouse lung to inhaled ozone. Toxicol Sci, 107: 535

- 543. http://dx.doi.org/10.1093/toxsci/kfn253

596419

Van Winkle, L. S.; Baker, G. L.; Chan, J. K.; Schelegle, E. S.; Plopper, C. G. (2010). Airway mast cells in a rhesus model of childhood allergic airways disease. Toxicol Sci, 116: 313

- 322. http://dx.doi.org/10.1093/toxsci/kfq119 670301

Veninga, T. S. (1967). Toxicity of ozone in comparison with ionizing radiation. Strahlentherapie, 134: 469 -477. 040746

Wang, T.

- N.; Ko, Y.

- C.; Chao, Y.

- Y.; Huang, C.

- C.; Lin, R.

- S. (1999). Association between indoor and outdoor air pollution and adolescent asthma from 1995 to 1996 in Taiwan. Environ Res, 81: 239

- 24 7. 008105

Wang, X. Y.; Hu, W.; Tong S (2009). Long

- term exposure to gaseous air pollutants and cardio

- respiratory mortality in Brisbane, Australia. Geospat Hea lth, 3: 257

- 63. http://www.geospatialhealth.unina.it/fulltext.php?ida=72. 199990

Wenten, M.; Gauderman, W. J.; Berhane, K.; Lin, P. C.; Peters, J.; Gilliland , F. D. (2009). Functional variants in the catalase and myeloperoxidase genes, ambient air pollution, and respiratory

- related school absences: An example of epistasis in gene

- environment interactions. Am J Epidemiol, 170: 1494

- 1501. http://dx.doi.org/10.10 93/aje/kwp310

597084

Wilhelm, M.; Ritz, B. (2005). Local variations in CO and particulate air pollution and adverse birth outcomes in Los Angeles County, California, USA. Environ Health Perspect, 113: 1212 -1221. 088668

Wilson, D. W.; Plopper, C. G.; Dungworth, D. L. (1984). The response of the macaque tracheobronc hial epithelium to acute ozone injury: a quantitative ultrastructural and autoradiographic study. Am J Pathol, 116: 193 -206. 040044

Witschi, H. (1991). Effect s of oxygen and ozone on mouse lung tumorigenesis. Exp Lung Res, 17: 473

- 483. 042509

Witschi, H.; Espiritu, I.; Pinkerton, K. E.; Murphy, K.; Maronpot, R. R. (1999). Ozone carcinogenesis revisited. Toxicol Sci, 52: 162 -167. 011602

Witschi, H.; Wilson, D. W.; Plopper, C. G. (1993). Modulation of N

- nitrosodiethylamin e- induced hamster lung tumors by ozone. Toxicology, 77: 193 -202. 043206

Wollmann, H. A. (1998). Intrauterine growth restriction: definition and etiology. Horm Res, 49: 1 -6. 193812

Wood, A. M.; Harrison, R. M.; Semple, S.; Ayres, J. G.; Stockley, R. A. (2009). Outdoor air pollution is associated

withdisease severit y in a1

- antitrypsin deficiency. Eur Respir J, 34: 346

- 353. 597085

Woodruff, T. J.; Darrow, L. A.; Parker, J. D. (2008). Air pollution and postneonatal infant mortality in the United States,

1999

- 2002. Environ Health Perspect, 116: 110 -115. 098386

Wu, R.; Zhao, Y. H.; Plopper, C. G.; Chang, M. M.; (1999). Differential expression of stress proteins in nonhuman primate lung and conducting airway after ozone exposure. Am J Physiol, 277: L511

- L522. 012089

Yang, C.

- Y.; Hsieh , H.

- J.; Tsai, S.

- S.; Wu, T.

- N.; Chiu, H.

- F. (2006). Correlation between air pollution and postneonatal mortality in a subtropical city: Taipei, Taiwan. J Toxicol Environ Health A, 69: 2033 -2040. 090760

Zelac, R. E.; Cromroy, H. L.; Bolch, W. E. Jr; Dunavant, B. G.; Bevis, H. A. (1971). Inhaled ozone as a mutagen: I chromosome aberrations induced in Chinese hamster lymphocytes. Environ Res, 4: 262

- 282. 039743

March 20 11 7

- 75 DRAFT – DO NOT CITE OR QUOTEZelac, R. E.; Cromroy, H. L.; Bolch, W. E. Jr; Dunavant, B. G.; Bevis, H. A. (1971). Inhaled ozone as a mutagen: II effect on the frequency of chromosome aberr ations observed in irradiated Chinese hamsters. Environ Res, 4: 325 -342.

039752

March 20 11 7

- 76 DRAFT – DO NOT CITE OR QUOTE